Medical mineralogy and geochemistry

MEDICAL MINERALOGY and GEOCHEMISTRY 64 Reviews in Mineralogy and Geochemistry 64 FROM THE SERIES EDITOR The review c...

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MEDICAL MINERALOGY and GEOCHEMISTRY 64

Reviews in Mineralogy and Geochemistry

64

FROM THE SERIES EDITOR The review chapters in this volume were the basis for a two day short course on Medical Mineralogy and Geochemistry held at the United States Geological Survey in Menlo Park, California, U.S.A. (December 9-10, 2006) prior to the American Geophysical Union Meeting in San Francisco, California. This meeting and volume were sponsored by the Mineralogical Society of America, Geochemical Society, and the United States Department of Energy. Any supplemental material and errata (if any) can be found at the MSA website www.minsocam.org. Jodi J. Rosso, Series Editor West Richland, Washington November 2006

SHORT COURSE SERIES DEDICATION Dr. William C. Luth has had a long and distinguished career in research, education and in the government. He was a leader in experimental petrology and in training graduate students at Stanford University. His efforts at Sandia National Laboratory and at the Department of Energy’s headquarters resulted in the initiation and long-term support of many of the cutting edge research projects whose results form the foundations of these short courses. Bill’s broad interest in understanding fundamental geochemical processes and their applications to national problems is a continuous thread through both his university and government career. He retired in 1996, but his efforts to foster excellent basic research, and to promote the development of advanced analytical capabilities gave a unique focus to the basic research portfolio in Geosciences at the Department of Energy. He has been, and continues to be, a friend and mentor to many of us. It is appropriate to celebrate his career in education and government service with this series of courses.

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DOI: 10.2138/rmg.2006.64.0

MEDICAL MINERALOGY and GEOCHEMISTRY 64

Reviews in Mineralogy and Geochemistry

64

FROM THE SERIES EDITOR The review chapters in this volume were the basis for a two day short course on Medical Mineralogy and Geochemistry held at the United States Geological Survey in Menlo Park, California, U.S.A. (December 9-10, 2006) prior to the American Geophysical Union Meeting in San Francisco, California. This meeting and volume were sponsored by the Mineralogical Society of America, Geochemical Society, and the United States Department of Energy. Any supplemental material and errata (if any) can be found at the MSA website www.minsocam.org. Jodi J. Rosso, Series Editor West Richland, Washington November 2006

SHORT COURSE SERIES DEDICATION Dr. William C. Luth has had a long and distinguished career in research, education and in the government. He was a leader in experimental petrology and in training graduate students at Stanford University. His efforts at Sandia National Laboratory and at the Department of Energy’s headquarters resulted in the initiation and long-term support of many of the cutting edge research projects whose results form the foundations of these short courses. Bill’s broad interest in understanding fundamental geochemical processes and their applications to national problems is a continuous thread through both his university and government career. He retired in 1996, but his efforts to foster excellent basic research, and to promote the development of advanced analytical capabilities gave a unique focus to the basic research portfolio in Geosciences at the Department of Energy. He has been, and continues to be, a friend and mentor to many of us. It is appropriate to celebrate his career in education and government service with this series of courses.

1529-6466/06/0064-0000$05.00

DOI: 10.2138/rmg.2006.64.0

1

Reviews in Mineralogy & Geochemistry Vol. 64, pp. 1-4, 2006 Copyright © Mineralogical Society of America

The Emergent Field of Medical Mineralogy and Geochemistry Nita Sahai Department of Geology and Geophysics University of Wisconsin Madison, Wisconsin, 53706, U.S.A. e-mail: [email protected]

Martin A. A. Schoonen Department of Geosciences Stony Brook University Stony Brook, New York, 11794-2100, U.S.A. e-mail: [email protected]

H. Catherine W. Skinner Department of Geology and Geophysics Yale University New Haven, Connecticut, 06520-8109, U.S.A. e-mail: [email protected]

Medical Mineralogy and Geochemistry is an emergent, highly interdisciplinary field of study. The disciplines of mineralogy and geochemistry are integral components of crossdisciplinary investigations that aim to understand the interactions between geomaterials and humans as well as the normal and pathological formation of inorganic solid precipitates in vivo. Research strategies and methods include but are not limited to: stability and solubility studies of earth materials and biomaterials in biofluids or their proxies (i.e., equilibrium thermodynamic studies), kinetic studies of pertinent reactions under conditions relevant to the human body, molecular modeling studies, and geospatial and statistical studies aimed at evaluating environmental factors as causes for activating certain chronic diseases in genetically predisposed individuals or populations. Examples presented in this volume include the effects of inhaled dust particles in the lung (Huang et al. 2006; Schoonen et al. 2006), biomineralization of bones and teeth (Glimcher et al. 2006), the formation of kidneystones, the calcification of arteries, the speciation exposure pathways and pathological effects of heavy metal contaminants (Reeder et al. 2006; Plumlee et al. 2006), the transport and fate of prions and pathological viruses in the environment (Schramm et al. 2006), the possible environmental-genetic link in the occurrence of neurodegenerative diseases (Perl and Moalem 2006), the design of biocompatible, bioactive ceramics for use as orthopaedic and dental implants and related tissue engineering applications (Cerruti and Sahai 2006) and the use of oxide-encapsulated living cells for the development of biosensors (Livage and Coradin 2006). Despite its importance, the area of Medical Mineralogy and Geochemistry has received limited attention by scientists, administrators, and the public. The objectives of the present short-course and volume are to highlight some of the existing research opportunities and challenges, and to invigorate exchange of ideas between mineralogists and geochemists 1529-6466/06/0064-0001$05.00

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working on medical problems and medical scientists working on problems involving geomaterials and biominerals. It is important to stress that universal physico-chemical principles fundamentally control the complex interactions of solutions, biomolecules, and minerals in the realms of the human body as well as in the natural geological environment (Frausto da Silva and Williams 1991; Williams and Frausto da Silva 1996; Sahai 2003, 2005). In both contexts, the molecular conformations, stabilities and reactivities of (bio)chemical components are critical to understanding specific reaction pathways. For example, silicate glasses and minerals such as Bioglass®, silica gel, wollastonite, pseudowollastonite, diopside, etc. have been shown to bond by the formation of a layer of carbonated hydroxylapatite at the implant/bone interface, making them suitable for use as orthopaedic and dental implant materials (Hench et al. 1971; Kokubo et al. 1982, 1994; Miake et al. 1995; de Aza et al. 1996; Liu and Ding 2001). An early step within this complex series of processes is the dissolution of the silicates by human blood plasma, involving reactions that are very similar to silicate weathering in the environment (Sahai and Anseau 2005; Cerruti and Sahai 2006). At the same time, it is critical to appreciate that biogeochemical reactions in vivo are ultimately under cellular and genetic controls, that acellular inorganic experiments or even in vitro experiments that include cells cannot entirely account for some of the unexpected, complex biological interactions, and that biominerals, because they are commonly nanocomposite materials, differ significantly in their physicalchemical and biological stability and reactivity compared to their bulk, macroscopic counterparts formed geologically or synthetically (Glimcher et al. 2006). Studies of the interaction between the human body and geomaterials have been conducted for some time but the advance in analytical techniques holds the promise of major breakthroughs over the next decade. More than six decades ago, V.M. Goldschmidt, the father of Geochemistry, contributed to a study of the effect of olivine on the lungs of rats (King et al. 1945). This study, led by Early Judson King, a prominent pathologist, was one of a series of studies directed at understanding silicosis in foundry workers. Since Goldschmidt’s early work, a number of geochemists and mineralogist have made contributions to the medical sciences, but progress has been hampered by the lack of insight at the molecular scale. The potential for uncovering the fundamental reaction mechanisms lies at this molecular level, leading to a deeper understanding of the overall process, and should increase our ability to design new treatment strategies. Recent developments in spectroscopy, molecular modeling, and high-resolution imaging provide powerful new avenues to explore the interaction between minerals and the human body at scales that are well beyond those accessible to Goldschmidt, his colleagues, as well as those that have been working in this area for the last few decades. While new techniques may enable advance, real progress will be made when geochemists, biomineralogists and biomedical researchers collaborate and harness the expertise in these clearly related disciplines. Geochemists and mineralogists are uniquely trained to contribute of this new field because of their knowledge of mineral stability, mineral reactivity, mineral precipitation/dissolution kinetics, and mineral-sorbate interactions, as well as their ability to study complex systems. While this short course focuses on biogeochemical interactions in vivo, other areas may be included within the broader umbrella of “Geology and Human Health” (Skinner and Berger 2003) such as observing and predicting natural disasters. Early warning systems for volcanic eruptions, earthquakes, hurricanes, land slides, tsunamis and other natural disasters are becoming globally available and could save millions of lives. Legends and historic records of such natural disasters as the massive volcanic eruptions that destroyed Pompeii and Herculaneum in 79 AD, the Laki eruptions in 1883-94 that killed almost a quarter of Iceland’s population with atmospheric disturbances over months that spread from Europe to Syria, and the eruption of Krakatoa, Indonesia, in 1883, underscore the importance of hazard

Medical Mineralogy and Geochemistry

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prediction. While most natural hazards are episodic, catastrophic, and deadly, some disasters, such as volcanic eruptions generating ash, may also present a long-term health burden that could take years to manifest itself as disorders and disease. Clearly, biogeochemical interactions in vivo are complex processes and understanding such processes at the molecular level will require a truly multi-disciplinary, integrated approach leveraging concepts and techniques from inorganic interfacial, aqueous and solid-state chemistry, computational chemistry, genetics, cellular and molecular biology. In addition to these scientific challenges, mineralogists, geochemists and biomedical scientists working in the field of Medical Mineralogy and Geochemistry field must meet the challenges of finding appropriate venues for research funding, peer-review, oral presentation of results and publication of research papers. In this regard, we are deeply appreciative of the forwardlooking approach by the Mineralogical Society of America and the Geochemical Society; the Environmental Geochemistry and Biogeochemistry Program, Division of Earth Sciences, National Science Foundation; and Chemical Sciences, Geosciences and Biosciences Division, Office of Basic Energy Sciences, Office of Science, U.S. Department of Energy, and the Crustal Imaging and Characterization Team, United States Geological Survey, in supporting and sponsoring this short-course and volume. In addition we are very grateful for the contributions of the authors as well as the efforts of the reviewers that helped improve the quality of this volume.

REFERENCES Cerruti M, Sahai N (2006) Silicate biomaterials for orthopaedic and dental implants. Rev Mineral Geochem 64:283-314 de Aza PN, Luklinska ZB, Anseau MR, Guitan F, de Aza S (1996) Morphological studies of pseudowollastonite for biomedical application. J Microscopy 182(1):24-31 Frausto da Silva JJR, Williams RJP (1991) The Biological Chemistry of the Elements: The Inorganic Chemistry of Life. Oxford University Press Glimcher MJ (2006) Bone: nature of the calcium phosphate crystals and cellular, structural, and physical chemical mechanisms in their formation. Rev Mineral Geochem 64:223-282 Hench LL, Wilson J (1993) An Introduction to Bioceramics. World Scientific Publishing CO Huang X, Gordon T, Rom WN, Finkelman RB (2006) Interaction of iron and calcium minerals in coals and their roles in coal dust-induced health and environmental problems. Rev Mineral Geochem 64:153-178 King EJ, Rogers N, Gilchrist M, Goldschmidt VM, Nagelschmidt G (1945) The effect of olivine on the lungs of rats. J Pathol Bacteriol 57:488-491 Kokubo T, Cho SB, Nakanishi K, Soga N, Ohtsuki C, Kitsugi T, Yamamuro T, Nakamura T (1994) Dependence of bone-like apatite formation on structure of silica gel. Bioceramics 7:49-54 Kokubo T, Shigematsu M, Nagashima Y, Tashiro M, Yamamuro T, Higashi S (1982) Apatite- and wollastonitecontaining glass-ceramics for prosthetic application. Bull Inst Chem Res Kyoto Univ 60:260-268 Liu X, Ding C (2001) Apatite formed on the surface of plasma sprayed wollastonite coating immersed in simulated body fluid. Biomaterials 22:2007-2012 Livage J, Coradin T (2006) Living cells in oxide glasses. Rev Mineral Geochem 64:315-332 Miake Y, Yanagisawa T, Yajima Y, Noma H, Yasui N, Nonami T (1995) High resolution and analytical electron microscopic studies of new crystals induced by a bioactive ceramic (diopside). J Den Res 74:1756-1763 Perl DP, Moalem S (2006) Aluminum, Alzheimer’s disease and the geospatial occurrence of similar disorders. Rev Mineral Geochem 64:115-134 Plumlee GS, Morman SA, Ziegler TL (2006) The toxicological geochemistry of earth materials: an overview of processes and the interdisciplinary methods used to understand them. Rev Mineral Geochem 64:5-57 Reeder RJ, Schoonen MAA, Lanzirotti A (2006) Metal speciation and its role in bioaccessibility and bioavailability. Rev Mineral Geochem 64:59-113 Sahai N (2003) The effects of Mg2+ and H+ on apatite nucleation at silica surfaces. Geochim Cosmochim Acta 67:1017-1030 Sahai N (2005) Modeling apatite nucleation in the human body and in the geochemical environment. Am J Sci 305:661-672 Schoonen MAA, Cohn CA, Roemer E, Laffers R, Simon SR, O’Riordan T (2006) Mineral-induced formation of reactive oxygen species. Rev Mineral Geochem 64:179-221

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Schramm PT, Johnson CJ, Mathews NE, McKenzie D, Aiken JM, Pedersen JA(2006) Potential role of soil in the transmission of prion disease. Rev Mineral Geochem 64:135-152 Skinner HCW, Berger AR (2003) Geology and Health: Closing the Gap. Oxford University Press Williams RJP, Frausto da Silva JJR (1996) The Natural Selection of the Chemical Elements. Oxford University Press

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The Toxicological Geochemistry of Earth Materials: An Overview of Processes and the Interdisciplinary Methods Used to Understand Them Geoffrey S. Plumlee U. S. Geological Survey Box 25046, Denver Federal Center MS 973 Lakewood, Colorado, 80225-0046, U.S.A. e-mail: [email protected]

Suzette A. Morman U. S. Geological Survey Box 25046, Denver Federal Center MS 964 Lakewood, Colorado, 80225-0046, U.S.A.

Thomas L. Ziegler Lakewood, Colorado, U.S.A. INTRODUCTION A broad spectrum of earth materials have been linked to, blamed for, and/or debated as sources for disease. In some cases, the links are clear. For example, excessive exposures to mineral dusts have long been recognized for their role in diseases such as: asbestosis, mesothelioma, and lung cancers (asbestos); silicosis and lung cancer (silica dusts); and coal-workers pneumoconiosis (coal dust). Lead poisoning, particularly in toddlers and young children, has been conclusively linked to involuntary ingestion of soils or other materials contaminated with lead-rich paint particles, leaded gasoline combustion byproducts, and some types of lead-rich mine wastes or smelter particulates. Waters with naturally elevated arsenic contents are common in many regions of the globe, and consumption of these waters has been documented as the source of arsenic-related diseases affecting thousands of people in south Asia and other regions. Exposure to dusts or soils containing pathogens has been documented as the cause of regionally common diseases such as valley fever (coccidioidomycosis) and much rarer diseases such as anthrax. Links between many other earth materials and specific diseases, although suspected, are less clear or are debated. For example, it has been suggested that geographic clusters of diseases such as leukemia are related to exposures to waters or atmospheric particulates containing organic or metal contaminants; however, for many clusters the exact causal relationships between disease and environmental exposure are difficult to prove conclusively. Even for many diseases in which the causal relationship is clear, such as in asbestosis and mesothelioma triggered by asbestos exposure, the minimum exposures needed to trigger disease, the influence of genetic factors, and the exact mechanisms of toxicity are still incompletely understood and are the focus of considerable debate within the public health community. Hence, understanding the health effects resulting from occupational and environmental exposures to a wide variety of earth materials remains a very active and fruitful area of research. As noted in previous papers (Fubini and Areán 1999; Plumlee and Ziegler 2003, 2006), the body interacts both physically and chemically with earth materials to which it is exposed 1529-6466/06/0064-0002$10.00

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(and which it may take up) via inhalation, ingestion, and dermal absorption. Although these interactions are ultimately physiological, they are also strongly influenced by the physical and chemical properties of the earth materials, coupled with their geochemical reactivities in the variety of water-based, chemically diverse body fluids (respiratory, gastrointestinal, perspiration, blood serum, interstitial, and intracellular fluids) encountered along the different exposure and uptake pathways. As a result, there are significant opportunities for earth scientists working in collaboration with health scientists to contribute expertise in mineralogy, materials characterization, fluid-mineral reactions, and other areas in the quest for a better understanding of the specific links between earth materials and human health. In order to emphasize the importance of such interdisciplinary approaches, Plumlee and Ziegler (2006) defined toxicological geochemistry (TG) as the study of the geochemical interactions between body fluids and earth materials, and how these interactions may influence toxicity. This chapter first provides an overview of the processes by which earth materials interact with the body. We then describe how earth science expertise and methods can be applied collaboratively throughout the range of heath science characterization and assessment methods commonly used to understand health issues associated with exposure to and uptake of earth materials. This paper is intended to complement and update the discussions presented in Plumlee and Ziegler (2003, 2006). As with these two earlier papers, we hope that readers from both the earth science and health science communities will find this chapter useful in understanding how geochemical processes and principles can be applied to help understand links between human health and exposures to certain earth materials. Our intent is to not imply that all earth materials are hazardous to health, as this is clearly not the case. Rather, we instead hope to show how a healthy dose of earth science expertise can be used to help understand where potential problems exist and help define the nature of the problems. We also hope that this paper and the many excellent other papers presented in this interdisciplinary volume will help identify and inspire future opportunities for fruitful research collaborations between geochemists, other earth scientists, and their counterparts in the health science community.

EARTH MATERIALS LINKED OR POTENTIALLY LINKED TO HUMAN HEALTH Earth materials can include a broad range of solid, gaseous, or liquid substances that are produced and released by natural earth processes, that are contaminated by and/or released from the earth as a result of human activities, or that are produced from the earth by humans and transformed for use in society (Fig. 1). Examples of earth materials with known or postulated health concerns are listed in Table 1. The health concerns have a variety of origins, and can result from the earth materials themselves (such as mineral dusts), from contaminants or pathogens contained within or carried by the earth materials, and from consumption of waters, plants, or organisms that have picked up contaminants by interacting with earth materials. The potential health concerns are quite diverse, and can include cancers, respiratory diseases, neurological diseases, diseases of the excretory system, secondary diseases such as congestive heart failure, increased susceptibility to pathogen infections, and many others. Historically, studies examining the toxicity and health effects of earth materials primarily focused on materials to which humans were exposed in workplace settings (such as asbestos or crystalline silica), or on soils or other earth materials containing toxicants of known health concern (such as lead, mercury, arsenic, or hexavalent chromium). Toxicological studies focused, for example, on commercially-produced asbestos or silica, or on the processes controlling toxicity of soluble metal salts or other materials common in workplace exposures. In the last several decades, there have been increasing concerns regarding environmental exposures to earth materials that a) are released into the environment by natural or anthropogenic processes, b) contain potential toxicants in more complex mineralogical forms than simple salts, or c)

Toxicological Geochemistry of Earth Materials

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Figure 1. There is a continuum of interactions between the earth’s geosphere, hydrosphere, biosphere, atmosphere, and anthropogenic activities that all can produce earth materials or environmental materials to which humans can be exposed and that may be of potential health concern.

contain other elements such as antimony, tungsten, selenium, vanadium, or others whose toxicity effects are less well-known.

FACTORS INFLUENCING THE HEALTH EFFECTS OF EARTH MATERIALS For detailed discussions on the toxicity effects of a wide variety of potential toxicants, the interested reader is referred to several excellent toxicology textbooks (such as Klassen 2001; Sullivan and Krieger 2001). Discussions focused on earth materials are presented by Plumlee and Ziegler (2003, 2006). There are many different factors that can influence the health effects of earth materials, including:

•

Intensity and duration of the exposure (the dose).

•

Exposure route.

•

Chemical conditions encountered along the exposure route.

•

Physical and chemical characteristics of the material.

•

The potential microbial or other pathogens present in the material.

Dusts from industrial, commercial asbestos products (insulation, brake linings, building products, others) and industrial activities. Dusts from asbestos accessory minerals in other industrial/commercial products (some, but not all, vermiculite, talc, and other products). Dusts from natural asbestos-bearing rocks and soils (via natural weathering and erosion or human disturbance) may provide low-level exposures that may be of concern to chronically exposed populations.

Dusts generated by mining, sandblasting, and other industrial or workplace activities. Dusts produced by erosion of friable, silica-rich rocks (i.e., some ash flow tuffs, diatomaceous earth deposits) may also play a role in disease. Dusts, other aerosols, and gases generated by coal mining, processing, and combustion activities.

These include talc, kaolinite, other clays, micas, and aluminosilicates.

Asbestos

Crystalline silica

Coal dust, coal fly ash, coal combustion gases

Other commercial and industrial mineral dusts Sources include dusts generated from industrial and commercial activities and products

Examples of potential exposure sources

Material

Skinner et al. (1988); Holland and Smith (2001); Pfau et al. (2005); Pan et al. (2005); Dodson and Hammar (2006); Roggli et al. (2006).

Inhalation.

Castranova and Vallyathan (2000); Castranova (2000); Daroowalla (2001);

Inhalation.

Mineral-specific fibrosis, such as talcosis; also silicosis and asbestosis due to intermixed crystalline silica and asbestos.

Inhalation.

Sulfur dioxide and volatilized arsenic released as a result of coal combustion have also been linked to health effects.

(table continued on next page)

Daroowalla (2001)

Castranova and Vallyathan (2000); Castranova (2000); Daroowalla (2001)

Silicosis, industrial bronchitis with airflow limitation, progressive massive fibrosis. Associated illnesses include opportunistic infections, silica nethropathy, lung cancer.

“Black lung disease” includes Coal Worker’s Pneumoconiosis (CWP), progressive massive fibrosis, chronic airway obstruction or bronchitis, emphysema; possible silicosis due to intermixed crystalline silica.

SSDC (1988);

Inhalation.

Has been proposed as a trigger of GI cancers, other health effects; however, links have not been demonstrated with certainty.

Ingestion.

Asbestosis; lung cancer; pleural effusion, thickening, plaques; mesothelioma cancer. Associated secondary illnesses include heart failure, other cardiovascular problems, lung infection, possible autoimmune responses.

References

Primary exposure routes Health effects

Table 1. Examples of earth materials and their sources with known or postulated health effects, modified from Plumlee and Ziegler (2003, 2006).

8 Plumlee, Morman, Ziegler

Examples of potential exposure sources Dusts generated from desert areas, dry lake beds, agricultural areas during dry periods.

Atmospheric particulates generated by petroleum or coal combustion, construction/demolition, industrial emissions, disturbance of urban soils, many others.

Glass wool (fiberglass); mineral wool (slag, rock wool)

Dusts from cement, concrete manufacturing, Concrete dusts generated by demolition, construction activities

Soils purposefully consumed by humans as a result of cultural practices or perceived nutritional needs.

Material

Natural mineral dusts

Urban particulates

Man-made mineral fibers.

Cement / concrete dust

Edible soils

Table 1. (continued)

Potential uptake of heavy metals or other toxicants. Secondary physiological problems associated with consumption of non-digestible mineral material.

Ingestion.

Irritation of eyes, throat, respiratory tract; ulceration of mucous surfaces. Effects largely tied to alkalinity of the dusts, although some heavy metals such as thallium or hexavalent chromium may be present. Silicosis maybe a concern with long-term exposure to concrete dust.

Inhalation, contact with exposed mucous membranes or moist skin.

Irritation of upper respiratory tract, skin. No ties to lung cancers, lung fibrosis.

Inhalation.

Possible uptake of heavy-metal and organic toxicants. Inhalation of acid aerosols of sulfur dioxide. Inhalation of irritant and/or oxidant gases such as nitric oxides and ozone. Increased asthma and cardiac stress, decreased lung function.

Inhalation.

Possible exposure to organic contaminants or pathogens in the dusts

Possible uptake of metals or metalloids such as arsenic.

(table continued on next page)

Abrahams et al. (2005)

Sahai (2001)

Hesterberg et al. (2001)

Costa (2001)

Derbyshire (2005)

Inhalation. Possible silicosis, asbestosis.

References


Toxicological Geochemistry of Earth Materials 9

Particulate matter (including smoke, ash, and fine soil particulates) and gases generated by forest fires.

Soils, and dusts generated from soils, that host pathogens such as bacteria and fungi .

Pathogens

Irritation of eyes, throat, respiratory tract; ulceration of mucous surfaces. Effects largely tied to acidity of the gases and droplets. If gases are in sufficiently high concentration, toxic effects can also result.

Acidic aerosol droplets formed when hot lava contacts seawater and causes it to boil.

Wildfire smoke, and ash

Inhalation, contact with exposed mucous membranes or moist skin.

Sulfur dioxide, hydrogen fluoride, hydrogen chloride, other acid gases emanating from active volcanoes.

Volcanic gases, vog, laze

Pathogen-related infections can develop through breaks in the skin.

Percutaneous absorption.

Asthma and pathogen-specific diseases such as Valley Fever (from the soil fungus C. Immitis) and anthrax (B. anthracis).

Inhalation, ingestion.

Asthma, irritation of respiratory tract; possible increased susceptibility to infection of respiratory system; possible long term increased cancer risk. Most studies have focused on organic constituents of smoke; less work has been done on mineral particles and heavy metals in smoke.

Inhalation.

Irritation of respiratory tract; asthma; potential effects of crystalline silica within ash.

Bultman et al. (2005)

Reinhardt and Otmar (2000); Reinhardt et al. (2000); Ward (1999); Wolfe et al. (2004)

Weinstein and Cook (2005), and their references.

Weinstein and Cook (2005), and their references.

Inhalation.

Atmospheric particulates generated by eruptions. Natural and anthropogenic disturbance of volcanic ash deposits, such as earthquakes, landslides, construction activities.

Volcanic ash

References


Examples of potential exposure sources

Material

Table 1. (continued)

10 Plumlee, Morman, Ziegler


11

•

Biosolubility, biodurability, bioaccessibility, and bioreactivity of the material in the body fluids encountered along the various exposure routes.

•

The body’s immune response.

•

The body’s physiological processes that control absorption, distribution, metabolism, and excretion of toxicants.

•

Other confounding factors, such as age, gender, genetics, personal habits (i.e., smoking), and personal socioeconomic, health, nutritional status, and dietary cofactors that may promote or counteract the toxic effects of the exposure.

Intensity and duration of exposure (the dose) To paraphrase an observation first made by Paracelsus in the 1500’s, “everything is a poison, nothing is a poison, it is the dose that makes the poison.” In its simplest definition, the dose is the amount of a potential toxicant that is taken up by an organism over a given period of time, with dose-response indicating that the greater the amount of a toxicant taken up by an organism, the greater the toxicological effect (Rozman and Klaasen 2001). However, toxicity is ultimately more complicated in that it is a function of the amounts of a potential toxicant that are taken up by the body, that survive the body’s many clearance and mitigation mechanisms, and that reach the particular site(s) of toxic action within the body. Each substance will have a threshold level above which it will become toxic that depends on the substance, the exposure route, and the individual (Sullivan et al. 2001). Substances can be acutely toxic, if the dose is sufficiently high over a short period of time, and/or chronically toxic, where the dose is lower but over a longer period of time. A wide variety of chemical elements, including major chemical species (sodium, calcium, potassium, etc.), and trace elements (zinc, copper, selenium, chromium, etc.) are essential to the effective functioning of the body, and so disease can result if these elements are deficient in the body; however, they can also become toxic if present in excess in the body (Lindh 2005a,b).

Exposure route There are several major exposure routes via which individuals can come into contact with potential toxicants or pathogens. These include inhalation, ingestion, and direct contact through unbroken skin (percutaneous), eyes (ocular), or wounds. A given toxicant or pathogen can have very different effects depending upon the exposure route due to the differences in physical processes, physiological processes, and chemical conditions to which it is subjected along each of the pathways (Fig. 2). Inhalation. A variety of physical, chemical, and physiological processes within the respiratory tract affect the ultimate disposition and health effects of inhaled earth materials (see overviews by Schlesinger 1988, 1995; McClellan and Henderson 1995; Fubini and Areán 1999; Gehr and Heyder 2000). Important disposition mechanisms include deposition, clearance, retention, and, for gases or soluble materials, absorption. The characteristics of an inhaled earth material play a complex but important role in its disposition, including: the type of material inhaled (solid, liquid, gas, or pathogen); the concentrations of the material in the inhaled air; the chemical composition of the material; the solubility and reactivity of the material in the fluids lining the respiratory tract; and, for solids, their shape, size distribution, density, and surface charge. Solid particles that deposit in approximately the first twelve generations of airways (which are lined by ciliated epithelium) are cleared by mucocilary clearance. Rhythmically-beating cilia waft the particle on a mucociliary escalator to the pharynx, where it is then swallowed. Particles depositing in more distal generations are subject to alveolar clearance, which is believed to occur primarily through phagocytosis by alveolar macrophages.

12


Saliva slightly acidic to slightly alkaline pH; high K, HCO3-, P; high enzymes; glycoproteins

Mucus mucoproteins, water Nasal Cavity

Oral Cavity Pharynx Larynx

Sweat acidic pH, high urea, lactic acid, etc.

Cilia

Trachea

Aqueous fluid near neutral pH, high PO2

Primary bronchi Secondary bronchi

Interstial fluids, plasma exposed at breaks in skin: pH 7.4; plasma has high proteins

Terminal bronchioles

Intracellular fluids pH 7.4; high K, Mg, P

Large intestine near neutral pH, reducing Stomach pH 1.5, electrolytes, enzymes

Interstitial fluids

Alveolar duct, sac, pH 7.4; high Na, low K; low proteins and alveoli

Small intestine near-neutral pH, reducing

Alveolar macrophage acidic lysosomal fluids with acid protease enzymes Capillary—plasma pH 7.4, moderately high PO2, high proteins

Lung fluid pH 7.4; high PO2; high Na, low K; mucoproteins, phospholipids

Figure 2. This schematic diagram, modified from Plumlee and Ziegler (2006), illustrates the various exposure routes and the substantial variability in body fluid types and compositions that can be encountered by earth materials during exposure.

The respiratory tract airways are lined with two liquid phases that facilitate deposition and clearance of solid particles: these include a mucous phase containing glycoproteins and phospholipid surfactants, and an underlying aqueous electrolyte phase. Mechanisms of particle deposition include (Fig. 3; Schlesinger 1988, 1995):

•

impaction on the airway walls at places where there are rapid changes in airflow direction (such as at branches or constrictions in the airway);

•

sedimentation onto the airway walls due to gravity settling;

•

electrostatic deposition of freshly generated particles with high surface charges;

•

interception of elongated particles by the airway walls;

•

deposition of fine particles (8 µm long, 10 µg/dL are considered elevated) are found, then investigations are typically undertaken to determine potential sources for the lead exposure; these source characterization studies routinely have involved measurements of bulk lead concentrations in the material(s) of interest, and have increasingly involved more detailed mineralogical characterization or chemical speciation studies to determine the form(s) of the lead. For example, Mielke (1999) and Mielke et al. (2004, 2006, and references therein) noted that 2030% of the children living in inner-city New Orleans prior to Hurricane Katrina had elevated blood lead levels and that there was a direct correlation between child blood lead and residence in census tracts with elevated soil lead. They also noted a direct correlation between elevated soil lead and elevated lead in interior house dusts. The source of the lead in the soils and blood is likely lead in paint (dispersed into the soils as a result of paint removal practices from building exteriors), and possibly remnant leaded gasoline combustion products remaining in the soils. There are a variety of shortcomings of BLL biomonitoring, such as the short half life (21-30 days) of lead in blood that precludes determination of non-recent exposures, and the remobilization of lead from the skeleton, which can obscure blood lead from recent exposures. These shortcomings are leading to the application of analytical methods from the earth sciences to examine records of past lead exposure recorded in the bones or teeth. For example, Arora et al. (2006) found that the spatial distribution of lead in teeth dentine reflected blood lead levels measured at birth and one year of age, using laser-ablation ICP-MS and confocal laser scanning microscopy; they concluded that the spatial distribution of lead in human primary teeth may be useful in obtaining temporal information on possible lead exposure during pre- and neonatal periods. Tracking sources of blood lead using earth science methods. Radiogenic lead isotope analysis methods were developed originally for earth science applications, but have now

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become a common tool in the biomonitoring toolkit. They are used quite widely and effectively to infer exposure sources for blood lead (Gulson et al. 1994, 1996a). They also can aid in the understanding of physiological processes by which lead is taken up, stored, and remobilized in the body (i.e., Gulson et al. 1996b). In vivo lead uptake studies. Whereas biomonitoring can provide information on past or ongoing lead exposures, it is also highly desirable to be able to anticipate potential lead uptake that may result from future or present but unrecognized exposures to lead in a variety of specific source materials. Due to the complexity of chemical conditions in the human body, in vivo tests are viewed by toxicologists as perhaps the most useful means to assess and predict potential absorption of lead by children from various source materials. Casteel et al. (2006) provided a very recent summary of in vivo lead uptake tests in which juvenile swine were fed various lead-contaminated soils and soluble reference material (lead acetate) twice a day for 15 days. The resulting lead uptake, measured as relative bioavailability (RBA, compared to the amount of lead uptake from soluble lead acetate reference material) was assessed though measurements of blood lead during the study course and analysis of lead in liver, kidney, and bone samples collected at sacrifice following the administration period. The soil materials from various mining- and smelting-related superfund sites were well characterized to understand the mineral phase, particle size, and matrix association of the lead in each of the samples. The study found that the RBA of lead from the soils is a strong function of the lead mineralogy (Fig. 12): lead that is tied up in cerussite (lead carbonate) and with manganese oxides is highly bioaccessible, whereas lead in galena (lead sulfide), anglesite (lead sulfate) and mixed lead-metal oxides and silicates, and iron sulfates, is relatively non-bioaccessible. Casteel et al. (2006) concluded that their data were not sufficiently well defined that RBA could be predicted based solely on the lead mineralogy and speciation. As discussed by Plumlee and Ziegler (2003, 2006), there are other factors in addition to the mineralogic host, particle size, and associated matrix that may influence biosolubility and bioaccessibility of lead and other metals, such as the texture of the mineral host, and the presence of absence of other trace elements in the mineral. In vitro bioaccessibility tests. Because in vivo studies are quite expensive, time-consuming, and logistically complicated, in vitro extraction tests using simulated body fluids (sometimes termed physiologically-based extraction tests, or PBET) have been used to model bioaccessibil-

1.0

0.8

Estimated lead relative bioaccessibility

0.6

0.4

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Le

ad -m Iro et a n m l ox et al ides su Le lfa ad tes su L Iro ead lfide n m sul Le eta fate lo ad ph xide os s ph L M an ea ates ga d o ne xid Le se o es ad xi ca des rb on at e

0.0

Figure 12. Plot from Casteel et al. (2006) showing the relatively bioavailability of lead from soils (determined by swine uptake studies) as a function of the lead-bearing minerals present in the soil.


47

ity of lead, other metals, and some organic toxicants in a variety of earth materials. Numerous studies have examined lead bioaccessibility from soils, mine wastes, and smelter byproducts by using a variety of different extraction methods, simulated gastrointestinal leachate fluid recipes, and solid:leachate ratios (i.e., Hamel et al. 1999; Ruby et al. 1999; Oomen et al. 2002, 2003, 2006; Shroder et al. 2004; Drexler and Brattin 2006; references in Plumlee and Ziegler 2003, 2006). Some of these tests model the entire passage through the gastrointestinal tract, by progressively subjecting a sample of the earth material of interest to simulated saliva for several minutes, followed by simulated gastric fluid for 1-2 hours, and then simulated intestinal (duodenal) fluids for 3-4 hours (i.e., Hamel et al 1999; Oomen et al. 2006). These tests can also be modified to simulate lead bioaccessibility and absorption in fasted versus fed conditions (with the addition of infant formula supplemented with sunflower oil), because metal bioaccessibility and absorption is generally greater in fasted than in fed conditions (Oomen et al. 2006). Shroder et al. (2004) and Drexler and Brattin (2006) summarize a simpler test that was developed in conjunction with, and best models results of, the in vivo juvenile swine study discussed in the previous section (Casteel et al. 2006). This test uses a simple simulated gastric leach with hydrochloric acid and glycine; key features are that the pH must be maintained near 1.5, and that the test is most specific for lead. Oomen et al. (2006) provide an excellent overview of the various tests, including a comparison of the saliva-gastric-intestinal tests versus gastric. For example, tests that only examine metal bioaccessibility in the gastric compartment overestimate total metal bioaccessibility, due to the diminished metal solubility and hence diminished bioaccessibility in the higher pH conditions of the intestinal tract. Also, soils with lead speciation and mineralogy not previously examined, as well as soils containing metals of interest other than lead, must ideally be calibrated against in vivo uptake tests to maximize the accuracy of the predicted bioaccessibility values. Although not specifically discussed by Oomen et al. (2006), it is possible that factors such as the amounts of iron-oxide particulates in the earth material may influence lead solubility and absorption in the intestinal tract by providing a substrate for lead sorption. Lead uptake models have been developed to help estimate potential risk from exposure to lead-rich sources at specific sites. Examples include the US EPA Integrated Exposure Uptake Biokinetic Model for Lead in Children (IEUBK) (EPA TRW 1999, and references therein), and the International Commission on Radiological Protection (ICRP) Pb model (see example of its use and references in Abrahams et al. 2006). These models predict blood lead levels in children that would result from consumption of lead-containing particles from a variety of sitespecific sources, including contaminated soils, lead paint chips, dust, foodstuffs, etc. Where data on these sources are not available, default values are assumed. The models also typically assume a default value for lead bioavailability (30% in the IEUBK model), which can be modified with the use of well-constrained bioaccessibility data developed through site-specific animal uptake or in vitro bioaccessibility testing (EPA TRW 1999). Figure 13 compares the total lead content of and amounts of lead leached from a variety of earth materials by a simple gastric leach (our unpublished data, obtained primarily following the protocols of Shroder et al. 2004, and Drexler et al. 2006). Compared are: two samples of lead-zinc mine wastes from the Kabwe, Zambia lead-zinc mine (samples provided by Dr. Gary Krieger); NIST (National Institute of Standards and Technology) Butte, Montana soil standard, affected by copper- and arsenic-rich mine wastes; two urban atmospheric particulate standards collected by the in the mid-1970’s in St. Louis, Missouri, and Washington, D.C.; a sample of the fine fraction of ballast lining a subway railbed; a NIST coal fly ash standard; two NIST soil standards with known contamination from lead paint chips; two edible soils purchased in a market in Kabwe (samples from Dr. Gary Krieger); a sample of ash left after a wildfire in ponderosa pine forest, Colorado front range (provided by Dr. Deborah Martin); and a sample of lake-bed dust from Owens Lake, California (provided by Dr. Marith Reheis). The

48

Plumlee, Morman, Ziegler 100 percent

10 percent

EPA playground bare soil screening level (400 ppm total lead )

Total Pb in solid Pb leached from solid by gastric leach

Lead Concentration

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Plumlee MSA Figure 13

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-ca

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ne wa

ste

with

Pb mo in F stly e-o rbo xide nate s* mine was te* N NIS IST TU Bute rba so n pa il NI i t r c ST Urb 1 64 an p 8 art Sub i c way b 16 4 al a 9 st f ine NIS T NI Co ST al f so ly as NI il with ST h lea ga dp rde aint n so il wi Pb pa int Ed ible so il Ed ible CO so il wildf Owe ire as ns d h ry lak ed ust

1 ppm

Figure 13. Our unpublished data on bulk metal concentrations and metal bioaccessibility from diverse earth materials indicate (in agreement with studies such as Casteel et al. 2006) that there can be substantial differences between the total lead concentration and lead bioaccessibility between different earth materials. Total metals were determined by ICP-MS following 4-acid digestion. Most of the leach measurements were made using the physiologically based gastric-fluid extraction test found to best reproduce lead uptake by swine (Casteel et al. 2006; Drexler and Brattin 2006); one part solid was added to 100 parts by weight simulated gastric fluid (HCl, glycine), and the mixture rotated at 37 °C for 2 h. Leach tests for samples indicated by the asterisk (*) were made using an NaCl-HCl simulated gastric fluid at 1 part solid to 40 parts gastric fluid. Lead concentrations in the leachates from both tests were recalculated to milligrams leached per kilogram solid for direct comparison to the total lead concentrations. In the NaCl leach tests, the pH shifted to substantially higher values (pH 4-5) due to the higher solid:liquid ratio and abundant carbonates in the samples; as a result, the amounts of lead leached should be an underestimate.

results show that Pb-Zn mine wastes (particularly those with high levels of lead carbonate), urban atmospheric particulates from the 1970’s, and soils containing lead paint chips have both high to very high lead contents, as well as high lead bioaccessibility; many samples have lead contents far above the 400 ppm EPA soil screening level for bare playground soil. The urban particulate standards have elevated lead levels and very high lead bioaccessibility due to the presence of combustion byproducts of leaded gas, which was banned in the 1970’s; hence most modern urban atmospheric particulates should have substantially lower lead levels than the NIST standards. The two mine waste samples illustrate the difference in lead bioaccessibility between lead largely tied up in iron oxides (less soluble in the gastric fluids) and lead tied up largely in lead carbonates (highly soluble in the gastric fluids). The African edible soil sample with elevated lead was purchased in a market near the now-closed lead-zinc mine from where the two Kabwe mine waste samples were obtained, and so it appears that the edible soil may have been produced from a soil contaminated by mine wastes or from a pre-mining soil developed on lead-mineralized rocks. The elevated lead levels and high bioaccessibility of the lead are obviously a source of health concern, as the soil is consumed mostly by pregnant women at a rate of tens to hundreds of grams per day; this could translate into the consumption of many tens to hundreds of milligrams of bioaccessible lead per day.

49


Lead uptake from earth materials via inhalation exposure? While many tests have been carried out to investigate the potential uptake of lead from earth materials by ingestion, far fewer have investigated potential lead uptake from earth materials via inhalation exposure (see summary in Plumlee and Ziegler 2003, 2006). We have investigated biosolubility of and metal bioaccessibility from a number of different earth materials using simulated lung and serum-based fluids, and have rather consistently found that lead does not show appreciable bioaccessibility in these simulated fluids (see the example of Owens Lake dusts shown in Fig. 14). In part, this may result from the lesser extent to which lead-bearing phases are solubilized in these pH 7.4 fluids (in spite of the abundance of myriad organic ligands available to complex the lead). However, it also likely results from the reaction of any solubilized lead with phosphate in the fluids to precipitate an insoluble lead phosphate phase. In contrast to lead, however, these tests suggest that oxyanion-forming (and potentially toxic) metalloids 10

35

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Figure 14. Our unpublished data comparing results of leach tests using simulated lung, gastric, and serumbased leach fluids of dust samples collected near Owens dry lake, CA, show that oxyanion species such as arsenic may be effectively leached by near-neutral pH fluids lining the lungs, due to their enhanced chemical mobility at near-neutral pH values. In contrast, lead is typically not leached by the simulated lung and serum-based fluids, due primarily to lead’s limited solubility at pH 7.4 and its reaction with phosphate in the fluids to form insoluble lead-phospate precipitates. Lung and serum-based fluid compositions are listed in Plumlee and Ziegler (2003, 2006). All simulated lung fluid tests were run using a 1:40 solid:liquid ratio for 24 h at 37 °C, with continuous rotation of the sample container.

50


such as arsenic, chromium, molybdenum, tungsten and antimony (not shown) are relatively bioaccessible in the simulated lung and serum-based fluids (Fig. 14), due to the enhanced mobility of the oxyanion species at pH 7.4. We are currently evaluating metal bioaccessibility from earth materials in acidic simulated lysosomal fluids, as a simplified proxy for potential metal uptake via particle phaogcytosis.

SUMMARY In this chapter, we have provided a comparatively brief overview of the myriad potential geochemical and biochemical processes that can occur when earth materials come into contact with body fluids via inhalation, ingestion, or percutaneous exposure routes. We have also shown how a wide variety of earth science methods have been and can be integrated with health science methods to better understand the potential health effects that might be associated with exposure to diverse earth materials. As noted by Plumlee and Ziegler (2003, 2006), it is possible to group individual earth material components according to how they behave in vivo, as well as by similarities in how they trigger toxicity:

•

Bioreactive earth materials can substantially modify the chemical composition of body fluids and tissues, to produce tissue irritation or more serious alkali or acid burns.

•

Solubilization of bioaccessible toxicants from earth materials and their subsequent absorption can produce toxic effects in the body.

•

Exposure to earth materials that are insoluble (biodurable) in body fluids can trigger toxic responses as the body attempts to clear the materials.

•

Pathogens associated with earth materials can trigger disease.

•

The body’s immune response to the earth materials or toxicants contained within the earth materials can trigger a toxic response.

All of these different these types of health effects are strongly influenced by the forms in which earth materials are delivered to the body (mineralogy; particle size and morphology; particle solubility, alkalinity, acidity; oxidation state of contained constituents; and others). These health effects all also ultimately require some level of chemical interactions between the earth materials and body fluids. However, most earth materials are complex mixtures of many different components, each having a particular mix of bioreactivity, biosolubility, and biodurability characteristics. Potentially complex chemical interactions between the various earth material components and the body’s fluids may result. Hence, as the potential health effects of complex earth materials are assessed, it should always be kept in mind that the integrated physical and chemical characteristics, chemical behavior in vivo, and resulting toxicity effects of the whole material may be substantially different from those of its individual components.. These conclusions implicitly require important roles for the earth scientist in helping to both characterize earth materials and understand geochemical processes in the context of the physiological processes of the human body. Many studies have been carried out to address this complex but fascinating topic; however, as indicated by the sections on asbestos and lead, there are many unresolved questions remaining. There should therefore be ample opportunities for fruitful future collaborations to occur between geochemists, toxicologists, physiologists, epidemiologists, and other biomedical disciplines.


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Todorov TI, Ejnik JW, Mullick FG, Centeno JA (2006) Chemical and histological assessment of depleted uranium in tissues and biological samples. In: Depleted uranium: properties, uses, and health consequences, AC Miller, ed, Lewis publishers (in press) UC Davis (2006) On the absestos trail. Synthesis, Fall-Winter 2006, UC Davis Cancer Center, accessed 8/2006 at http://www.ucdmc.ucdavis.edu/synthesis/Archives/fall_winter_06/features/outreach.html Van Gosen BS (2005) Reported historic asbestos mines, historic asbestos prospects, and natural asbestos occurrences in the Eastern United States. U.S. Geol. Survey Open-File Report 2005-1189, http:// pubs.usgs.gov/of/2005/1189/ Van Gosen BS (2006a) Using the geology of asbestos deposits to predict the presence or absence of asbestos in mining and natural environment. In: Proceedings of the 42nd Forum on the Geology of Industrial Minerals, Asheville, North Carolina, May 7-13, 2006. JC Reid (ed) North Carolina Geological Survey Information Circular 34 [CD-ROM], 412-432 Van Gosen BS (2006b) Reported historic asbestos prospects and natural asbestos occurrences in the Central United States. U.S. Geol. Survey Open-File Report 2006-1211, http://pubs.usgs.gov/of/2006/1211/ Van Gosen BS, Lowers HA, Sutley SJ, Gent CA (2004) Asbestos-bearing talc deposits, southern Death Valley region, California. In: Betting on Industrial Minerals—Proceedings of the 39th Forum on the Geology of Industrial Minerals, Reno/Sparks, Nevada, May 18-24, 2003. Castor SB, Papke KG, Meeuwig RO (eds) Nevada Bureau of Mines and Geology Special Publication 33:215-223 Vutchkov M, Lalor G, Macko S (2005) Inorganic and organic geochemistry techniques. In: Essentials of Medical Geology. Selinus O, Alloway B, Centeno J, Finkelman R, Fuge R, Lindh U, Smedley P (eds) Elsevier Inc., p 695-723 van Oss CJ, Naim JO, Costanzo PM, Giese RF Jr, Wu W, Sorling AF (1999) Impact of different asbestos species and other mineral particles on pulmonary pathogenesis. Clays Clay Min 47:697–707 WHO (1996) Trace Elements in Human Nutrition and Health. World Health Organization in collaboration with the Food and Agricultural Organization of the United nations and the International Atomic Energy Agency, 343 pp Walker JS (1981) Asbestos and the Asbestiform Habit of Minerals. Unpub. M.S. thesis, University of Minnesota, 175 pp Weinstein P, Cook A (2005) Volcanic emissions. In: Essentials of Medical Geology. Selinus O, Alloway B, Centeno J, Finkelman R, Fuge R, Lindh U, Smedley P (eds) Elsevier Inc., p 563-594 Werner AJ, Hochella MF Jr, Guthrie GD Jr, Hardy JA, Aust AE, Rimstidt JD (1995) Asbestiform riebeckite (crocidolite) dissolution in the presence of Fe chelators: implications for mineral-induced disease. Am Mineral 80:1093–1103 Wolfe MI, Mott JA, Voorhees RE, Sewell CM, Paschal D, Wood CM, McKinney PE, Redd S (2004) Assessment of urinary metals following exposure to a large vegetative fire, New Mexico, 2000. J Expo Anal Environ Epidemiol 14:120-128 Ward DE (1999) Smoke from wildland fires. In: Health Guidelines for Vegetation Fire Events, Lima, Peru, 6-9 October 1998, Background papers. World Health Organization, p 70-85 Wood S, Taunton A, Normand C, Gunter M (2006) Mineral-fluid interaction in the lungs: Insights from reactionpath modeling. Inhal Toxicol 18:975-984 Wright RS, Abraham JL, Harber P, Burnett BR, Morris P, West P (2002) Fatal asbestosis 50 years after brief high intensity exposure in a vermiculite expansion plant. Am J Resp Crit Care Med 165:1–5 Wylie AG, Verkouteren JR (2000) Amphibole asbestos from Libby, Montana, aspects of nomenclature. Am Mineral 85:1540-1542 Yang H, Bocchetta M, Kroczynska B, Elmishard AG, Chen Y, Liu Z, Bubici C, Mossman BT, Pass HL, Testa JR, Franzoso G, Carbone M (2006) TNF-α inhibits asbestos-induced cytotoxicity via a NF-κB-dependent pathway, a possible mechanism for asbestos-induced oncogenesis. Proc Nat Acad Sci 103:10397-10402 Ziegler JL (1993) Endemic Kaposi’s sarcoma in Africa and local volcanic soils. Lancet 342:1348-1351 Ziegler TL, Plumlee GS, Lamothe PJ, Meeker GP, Witten ML, Sutley SJ, Hinkley TK, Wilson SA, Hoefen TF, Brownfield IK, Lowers H (2002) Mineralogical, geochemical and toxicological variations of asbestos toxicological standards and amphibole samples from Libby, MT. In: Geol Soc Am Abst Progr, 2002 Annual Meeting, http://gsa.confex.com/gsa/htsearch.cgi

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Metal Speciation and Its Role in Bioaccessibility and Bioavailability Richard J. Reeder, Martin A. A. Schoonen Department of Geosciences and Center for Environmental Molecular Science Stony Brook University Stony Brook, New York, 11794-2100, U.S.A. e-mail: [email protected], [email protected]

Antonio Lanzirotti Consortium for Advanced Radiation Sources University of Chicago Chicago, Illinois, 60637, U.S.A. e-mail: [email protected]

INTRODUCTION Metals play important but varied roles in human health. Some metals are required for normal metabolic function, with optimal amounts for maximum benefit. Others are only known to cause toxic effects. Most of our knowledge of the function of metals in human health has been acquired in the last 100 years. However, evidence of adverse health effects attributed to metal exposures dates back to early civilizations. For example, it has been deduced that extensive mining and smelting of lead and its widespread use in the Roman Empire caused significant incidence of lead poisoning (Nriagu 1983; Hong et al. 1994). The source of metals in the environment ultimately can be traced back to their occurrence primarily in rocks, with their release to soil, water, and air facilitated by weathering processes. Consequently, the natural occurrences of metals in soils and waters are strongly correlated to the varied distribution of rock types and the compositions of the constituent minerals. More than 2000 years ago the Greek physician Hippocrates recognized relationships between disease and location, illustrating that environmental factors influenced human health. Today there are many known geographic patterns of disease that have been correlated with properties of soils or waters, or even aerosol particles. It has been difficult to demonstrate cause-effect relationships for many correlations, and efforts to relate total concentration of a metal contaminant to toxic impact have proven difficult (Davies et al. 2005). This point illustrates the essential concept that the total amount of an element in an environmental setting is not necessarily a good measure of its potential health threat. Within the last two centuries human activities have been highly effective in redistributing metals on local, regional, and even global scales. This has contributed to a greater exposure to humans. We have also changed the chemical forms of metals, sometimes with unfortunate consequences that include enhanced mobility in the environment as well as creation or enhancement of more toxic forms. Nearly three-quarters of the elements in the periodic table are classified as metals. Inasmuch as all but a few of them occur in nature, it is probably correct to say that each one has (or will be found to have) an important role with regard to environmental health. In this paper, we focus primarily on a few of the so-called heavy metals that are known to be 1529-6466/06/0064-0003$10.00

DOI: 10.2138/rmg.2006.64.3

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associated with adverse health effects. In a broad sense the heavy metals can be considered to include metals with atomic number greater than 20 (Ca), but many readers will know that a consistent definition has not found universal acceptance (Duffus 2002). Heavy metals should not necessarily be equated with toxicity, however, as a number of them are essential for human health in small amounts. Heavy metals have widespread occurrence as trace or minor constituents of soils, rocks, and waters, as well as living organisms (including humans). Commonly included among the heavy metals are arsenic, which is a metalloid, and selenium, a non-metal. Depending on factors such as oxidation state, electron configuration, ionic radius, and the presence of various ligands, metals exhibit a rich variety of coordination compounds in aqueous, solid, and even gaseous forms. The fact that the particular chemical form of a metal strongly influences its chemical behavior, mobility in the environment, uptake by organisms, and toxicity is certainly justification for characterizing and understanding metal speciation, not only in the environment but also in the body.

METAL SPECIATION CONCEPTS Speciation refers to aspects of the chemical and physical form of an element. Oxidation state, stoichiometry, coordination (including the number and type of ligands), and physical state or association with other phases all contribute to define speciation. These properties govern the chemical behavior of elements, whether in environmental settings or in human organs, and play a crucial role in determining toxicity. The focus on metal speciation in this chapter reflects the varied roles it plays in human health. Metals such as iron and zinc are essential for metabolic function, but can be toxic in excess. Others, like cadmium and lead, have no known beneficial function and pose health risks even at low levels of exposure and uptake. The amount of exposure or uptake is obviously a key factor in assessing adverse health impacts, and defines the field of toxicology. However, the metal speciation is also a critical factor in determining toxicity. For example, inorganic dissolved mercury (Hg2+(aq)) and methyl mercury chloride (CH3HgCl(aq)) are both considered to be toxic, but the properties and behavior of the latter make it a significantly greater health threat (NRC 2000). Another example is illustrated by the two common oxidation states of chromium in soils and water. Hexavalent chromium in the form CrO42− is soluble in water, making it mobile, and readily taken up by organisms. This form is also a known carcinogen (ATSDR 2000). Trivalent chromium tends to be insoluble, often forming hydroxide solids, and is considered an essential element in small amounts. We will see later that the pathways of uptake involve processes that may alter speciation, thereby changing a toxic form into a benign one, or the reverse. Although the concept of speciation is now widely appreciated in many fields, there have been few efforts to provide a formal definition. Bernhard et al. (1986) pointed out that usage varies among different fields, ranging from evolutionary changes to distinctions based on chemical state. A Molecular Environmental Science Workshop convened by the U.S. Department of Energy in 1995 cited at least five aspects important for defining speciation: element identity, physical state, oxidation state, chemical formula, and detailed molecular structure (DOE 1995). Both reports emphasized the importance of techniques available for determining these properties and the limitations they place in our ability to identify and distinguish species. Several Divisions within the International Union of Pure and Applied Chemistry (IUPAC) addressed speciation concepts in detail (Templeton et al. 2000), recommending that its usage in chemistry be restricted to distribution of chemically distinct species: “Chemical compounds that differ in isotopic composition, conformation, oxidation or electronic state, or in the nature of their complexed or covalently bonded substituents, can be regarded as distinct chemical species.” Although very similar to current usage in the Earth and environmental sciences, the IUPAC notion of speciation includes the isotopic identity of the element. Isotopic identify (i.e., mass differences within the nucleus) may have only a marginal effect on chemical

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behavior, and then only when mass differences are large, such as with light isotopes. However, isotopic identity could also be significant in instances where particular isotopes have special characteristics, such as radioactivity. The IUPAC concept of speciation does not draw attention to the significance of the physical state or association in distinguishing species. Earth and environmental scientists are particularly aware of the importance of physical associations. For example, occurrences of Cs+ as an aquo ion in a soil water solution, sorbed at the surface of a mineral, or exchanged into an interlayer site in a smectite all represent distinct species, even though they may share the same oxidation state, coordination, ligands, and other local properties. This last example also highlights the fact that multiple species of a metal commonly coexist in real systems. Where different species of a given metal share some property, such as oxidation state, it can be difficult to distinguish among them. In fact our ability to characterize metal speciation is dependent on the techniques used and the information they provide. For example, the use of a technique that gives direct information regarding oxidation state may fail to distinguish metal species having the same oxidation state but are complexed by different ligands. Identification of species using a mass spectrometry method may fail to distinguish oxidation state, but also may require sample processing that could alter speciation. Such problems become amplified in complex solutions such as lung or gastric fluids, which contain numerous organic components, including peptides, amino acids, and phospholipids. Hence, species assessment is operationally defined, and no single technique provides information concerning all relevant aspects that define a species. Advantages can often be gained by use of several complementary techniques for species characterization. In view of the fact that many of the metals most relevant for human health occur at very low concentrations, adequate characterization may pose serious challenges. Later is this chapter, we consider some of the more commonly used techniques for assessment of metal speciation.

SIGNIFICANCE OF SPECIATION The chemical and physical aspects that define speciation of a metal control its reactivity, including its solubility and uptake behavior, and, in many circumstances, toxicity. Solubility and uptake behavior, in turn, influence mobility of the metal in the environment, and therefore constrain pathways of exposure to organisms, including humans. During exposure the metal speciation directly influences absorption across a physiological membrane, which allows entry into systemic circulation. A transformation in speciation may occur in biological fluids (e.g., lung or gut fluids) prior to any absorption, however, which may affect absorption and subsequent toxicity. Within organ systems detoxification processes may further alter speciation and toxicity, and also influence transport, excretion, and storage. This oversimplified description illustrates the importance of metal speciation over the entire spectrum of process impacting the metal’s fate from weathering to human impact. Readers are referred to Plumlee and Ziegler (2003) and Plumlee et al. (2006) for more a comprehensive discussion of these aspects. The dependence of toxicity on speciation is now well known. The behavior of a metal may be completely changed by its oxidation state or its association with specific ligands, as exemplified by the contrasting toxicities of methylmercury and inorganic mercury species. The metalloid tin also shows markedly different health threats depending on its association with specific ligands. Neither metallic nor inorganic forms of tin present a health problem in small amounts; in fact, SnF2 is a common additive of toothpaste. However, many organotin compounds, which are predominantly created by human industrial processes, are highly toxic (ATSDR 2005c). Tributylin tin, widely used as a biocide and antifouling agent for seagoing vessels since the 1970s (de Mora 1996), is a potent ecotoxicant (Alzieu 1996; Maguire 1996), persists in marine environments (Diez et al. 2002; Sudaryanto et al. 2004, 2005), accumulates

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in tissue of fish and shellfish (Alzieu 1996; Sudaryanto et al. 2004), and may cause adverse health effects in humans (Kafer et al. 1992; Dopp et al. 2004). One of the complicating aspects of speciation is that each species exhibits a distinct behavior, making generalizations about stability and reactivity difficult. In this chapter, we do not attempt to provide a comprehensive review of metal speciation. Instead, our approach is to illustrate important aspects of metal speciation on human health using selected examples. As noted already, the total concentration of a particular element in any system, environmental or human, is not necessarily a good indicator of its potential health impact. Although this concept has been widely embraced by the research community and acknowledged by regulatory agencies, its impact on development of regulatory standards in the US has been limited. Even in the toxicology field, many bioassays do not consider speciation of metals. It is noteworthy, for example, that current methods for analysis of metals in soils (e.g., EPA 3050b) are designed to recover the more soluble metal fraction by use of acid digestion. As speciation techniques become more widely used and as understanding of the differences in behavior among species, including transformations, improves, it is likely that agencies will take greater account of these factors in formulating regulations. This is a critical area of research to which geochemists and mineralogist will be able to make important contributions.

ROLE OF METALS IN HUMAN HEALTH AND METAL TOXICITY CONCEPTS The human body requires the uptake of several essential metals for its proper function. As briefly summarized in Appendix 1, a number of heavy metals are known to be essential. The roles of some of the essential heavy metals listed in Appendix 1, such as vanadium and tungsten, are not fully known. For others, including arsenic and tin, essential roles have been suggested, but not demonstrated. Some metals play a role in active centers of metalloenzymes. In fact, for metals such as cobalt this may represent the dominant “species” in the human body. Other metals, such as chromium(III) and vanadium, are metabolized within the body to form low molecular weight compounds that play a role in glucose metabolism. The dose-response curve for an essential metal, schematically shown in Figure 1a, has a characteristic optimal range, flanked by suboptimal ranges. For some metals, a deficiency may be expressed as a specific disease. For example, chromium is important in the human metabolic system. A lack of chromium(III) disrupts glucose metabolism and can lead to obesity, diabetes, and cardiovascular disease, as well impairment of the reproductive system (Appendix 1).

b Mortality (percent)

Toxicity

Optimal

Deficiency

Function

a 100

LOAEL 50

NOAEL

0

Dose

0

25

50

Dose (mg)

Figure 1. Schematic dose-response curves for (a) an essential metal and (b) a non-essential toxic metal. NOAEL and LOAEL are no observed adverse effect level and lowest observed adverse effect level (see Table 1).


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It is important to note that deficiency in a metal may be caused by several factors. A diet lacking an essential metal is a common cause. For example, in China outbreaks of Keshan disease, a type of heart disease, are restricted to well defined geographic regions (Fordyce 2005). The occurrence of this disease is associated with a lack of dietary selenium. It is within this context that speciation is important. Crops grown on soil must be able to extract the selenium from the soils. Within Zhangjiakou District, Hebei Province, China, Keshan Disease has a high prevalence despite the fact that there is significant total selenium in the soil. In fact, a study in this region showed that the prevalence of the disease is not correlated with a lack of selenium in the soil as might be expected (Johnson et al. 2000). Rather, the cause for the selenium deficiency in the diet is a result of the fact that the soil-bound selenium is not in a form that is available to the plants. Soils in affected areas are rich in organic matter, and it is hypothesized that selenium is strongly bound with the organic fraction (Johnson et al. 2000). Alternatively, the organic matter may promote reduction of selenate to selenite, the latter commonly being more strongly sorbed to iron hydroxides in soils (Hartikainen 2005). This example illustrates the importance of speciation—in this instance in soils—rather than total concentration in understanding the incidence of a disease. While geological factors have been shown to contribute to diseases, such as Keshan Disease, additional confounding factors can lead to complex patterns in the prevalence of a disease. The interaction between metals (or metalloids) is one of the confounding factors. Nutritional status is known to affect toxicity. For example, anemia—low iron status—promotes the uptake of nickel and manganese. Hence, individuals affected by anemia are at a higher risk for adverse effects of nickel and manganese uptake compared to healthy individuals exposed to the same nickel and manganese levels. Conversely, exposure to methylmercury has been shown to inhibit uptake of selenium, an antioxidant (Norling et al. 2004). A second major confounding factor is genetic predisposition. Genetic disorders that disrupt the uptake or transport of essential metals or formation of antioxidants are more difficult to diagnose and remedy. For example, Hallervorden-Spatz syndrome (HSS) is a neurodegenerative disease caused by a genetic defect that disrupts the function of ferritin, an iron storage protein. HSS leads to accumulation of iron in the brain and is thought to cause oxidative stress (Zhou et al. 2001). Exposure and uptake of toxic metals, or even essential metals at levels in excess of the optimal range, can lead to adverse health effects. Exposure history is an important factor in evaluating the toxicity of a metal. Toxicologists distinguish between acute and chronic toxicity. Acute toxicity is that associated with short-term exposure to a toxicant, sometimes in lethal doses. Chronic toxicity is that associated with long-term exposure, and is usually the type most relevant to environmental toxicants. Studies based on rats or other laboratory animal models are commonly used to assess adverse health effects and to establish the no observed adverse effect level (NOAEL) and the lowest observed adverse effect level (LOAEL). A schematic doseresponse curve for a toxic metal or substance is shown in Figure 1b. The outcomes of these studies form the basis for regulations. An overview of the methodology behind such studies is given by the U.S. EPA (http://www.epa.gov). Table 1 defines common terms used in such studies. Toxicity data for metals and other susbstances are available (on-line) from the Agency for Toxic Substance and Disease Registry (ATSDR) within the U.S. Department of Health and Human Services (http://www.atsdr.cdc.gov). For a web-based introduction to the basic concepts in toxicology and its terminology the reader is referred to the National Library of Medicine (National Institutes of Health) web page: http://sis.nlm.nih.gov/enviro/toxtutor/Tox1/amenu.htm.

BIOAVAILABILITY AND BIOACCESSIBILITY CONCEPTS The term bioavailability is much used throughout the literature. There is a general understanding that the meaning addresses the potential for a substance to interact with an organism. With the use of this term now widespread among many disciplines, confusion

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Reeder, Schoonen, Lanziro i Table 1. Some abbreviations and terms relating to toxic effects of metals. Abbreviation

Definition

NOAEL

No-observed-adverse-effect level. The actual doses (levels of exposure) used in studies that showed no observable adverse effects to the organism.

LOAEL

Lowest-observed-adverse-effect level; LOAELS have been classified into “less serious” or “serious” effects. “Serious” effects are those that evoke failure in a biological system and can lead to morbidity or mortality (e.g., acute respiratory distress or death). “Less serious” effects are those that are not expected to cause significant dysfunction or death, or those whose significance to the organism is not entirely clear.

MRL

Minimal Risk Level; an estimate of daily human exposure to a substance that is likely to be without an appreciable risk of adverse effects (noncarcinogenic) over a specified duration of exposure

CEL

Cancer effect level

LD50

Lethal dose that leads to 50% mortality

Note: Other web resources that provide definitions of terms used in toxicology are http://extoxnet.orst.edu/ tibs/standard.htm and http://www.atsdr.cdc.gov/glossary.html.

sometimes develops as specific meanings emerge from a particular context, discipline, or method of study. An example is the meaning of bioavailability shared by the toxicology and pharmacology fields, where it refers specifically to the fraction of an administered dose that is absorbed into the organism’s circulatory system or into the organ where an effect occurs (Ruby et al. 1999; NRC 2003). The reason for such a restricted definition is clear upon consideration of the methodologies typically employed to evaluate efficiency of uptake of a drug or a toxicant. For example, a study might involve assays of blood levels of a given toxicant to identify peak plasma concentration and half-life resulting from specific oral dosages. Even this definition of bioavailability could find disfavor, since substances absorbed through the gastrointestinal tract of humans first circulate through the liver, where metabolism may limit the amount released to general systemic circulation. This view of bioavailability is of little value to the soil geochemist examining the fraction of a metal in a soil that becomes soluble and mobile during a sequential extraction procedure. In fact, there may be no universally acceptable definition of bioavailability, a prospect that may be most evident where operational definitions are desired. The lack of a common view of bioavailability was addressed by a recent report from the National Research Council (2003) in the context of soil and sediment contaminants, and summarized by Ehlers and Luthy (2003). We refer the reader to this publication for a more detailed description of various definitions of bioavailability and the rationale for their use. Rather than propose a working definition, the NRC study recommended the adoption of a process-based view, with bioavailability processes being defined as “the individual physical, chemical, and biological interactions that determine the exposure of plants and animals to chemicals associated with soils and sediments,” which is shown schematically in Figure 2. While a process-oriented view will be intuitive for many geochemists, the NRC perspective is purposely limited on the organism side to exclude processes following transport of a substance across the biological membrane. Because of the context relating to soil and sediment contaminants, it places emphasis on those factors and processes that make the substance “available” to the organism, that is, in a form that can be transported across the organism’s biological membrane. This is most commonly interpreted to be a soluble form. However, it is also possible that very small solids or colloidal particles could be transported across some


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Figure 2. Schematic diagram illustrating the range of bioavailability processes as defined in the NRC study (2003). [Reprinted with permission from Bioavailability of Contaminants in Soils and Sediments, Copyright (2003) by The National Academies of Sciences, Courtesy of the National Academies Press, Washington, D.C.]

membranes, such as the linings of lungs. Ruby et al. (1996, 1999) used the term bioaccessibility to represent the fraction of a toxicant (or substance) that becomes soluble within the gut or lungs and therefore available for absorption through a membrane. The amount actually absorbed, which according to the view described above reflects bioavailability, may be less than the soluble fraction. This concept of bioaccessibility carries over to the release or solubilization of soil- and sediment-associated metals in environmental systems external to the organism. Selective sequential extraction procedures are familiar examples of protocols for solubilizing certain constituents in geomaterials (e.g., Tessier et al. 1979; Scheckel et al. 2003, 2005). This concept of bioaccessibility also applies to in vitro studies that assess the release (solubility) of solid-bound metals in simulated biological fluids, as models of human lung, gastric, or intestinal fluid. Such physiologically based extraction tests (PBET) are being evaluated as in vitro alternatives to more costly in vivo studies using animal models (Ruby et al. 1996, 1999; EPA 2005). Plumlee and Ziegler (2003) describe bioaccessibility tests and comparisons with bioavailability for a variety of earth materials. Similar in vitro approaches have been widely used in pharmacological studies to assess drug release (e.g., Lin and Lu 1997). Agreement on the usage of bioavailability, bioaccessibility, and related terms seems unlikely in the near future, including within the environmental geochemistry and mineralogy fields. In response to the NRC report, Semple et al. (2004) suggested temporal and/or spatial distinctions between bioavailable and bioaccessible compounds. They proposed that the term bioavailable compound be used where the substance “is freely available to cross an organism’s cellular membrane from the medium the organism inhabits at a given time.” They proposed using bioaccessible to indicate a compound that “is available to cross an organism’s cellular membrane from the environment, if the organism has access to the chemical,” indicating that such a compound could subsequently become bioavailable once the proximity of the compound to the organism allows. Other authors have proposed the term geoavailability to represent the fraction of a toxicant or contaminant in a geologic material that becomes soluble or mobile as a result of various biogeochemical processes (Smith and Huyck 1999). This focus on terminology may appear to be only marginally relevant to the main topic. However, as Ehlers and Luthy (2003) note, bioavailability concepts may very well enter into risk assessment in the future and play an important regulatory role. Hence, we can expect further discussion of these concepts and related terminology, and it is incumbent on geoscientists to participate in this dialog.

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In this chapter we refrain from recommending a preferred usage, and merely alert the reader to the potential for confusion. However, having noted that bioaccessibility can be generalized to include solubilization of metals in geomaterials, we tend to follow that usage here and retain the distinction with bioavailability, which we will use to refer to absorption across a physiological membrane. As emphasized in the following section, we believe there is merit in taking a broad view of the role of metal speciation, encompassing processes from weathering and transport in geomedia, through exposure and uptake by humans, to fate within organs.

PATHWAYS OF METAL UPTAKE— FROM SOIL, WATER, AND AIR TO HUMAN ORGANS Historically, the gulf between the geosciences and the health sciences has tended to constrain the research activities in each community to address different parts of the broader subject of environmental health. Geoscientists have typically limited their attention to environmental processes that govern the bioaccessibility of metals up to and sometimes including exposure. Health scientists have naturally focused on the health impact, generally beginning with exposure. Clearly both communities recognize the continuum of processes that link the geologic and health aspects. In this chapter, the primary goal is to illustrate the role of metal speciation spanning both environmental and physiological processes. General exposure pathways are described by Plumlee et al. (2006) elsewhere in this volume (also see Plumlee and Ziegler 2003). Exposure pathways for metals are the same as for other toxicants (aka xenobiotics): ingestion (gastrointestinal tract), inhalation (respiratory tract), and dermal contact. A particular metal may be present in a solid, a liquid, or a gas phase. The physical form commonly dictates the nature of the exposure. For example, arsenic that desorbs from iron oxide-coated sands in an aquifer will enter groundwater that may be a source of drinking water. Here, ingestion of dissolved arsenic is the main exposure pathway. Ingestion is also possible for metals in solid forms, including food and soil particles. Certain associations of a metal may also dictate exposure. For example, inhalation would be the primary exposure pathway for arsenic associated with airborne particles (e.g., in mineral dust). If a fraction of inhaled particles is transported by mucociliary action to the pharynx, then ingestion may become a secondary exposure pathway.

Environmental processes Geochemists generally consider that metals are released to the human environment initially by weathering of rocks and are subsequently modified by various processes operating at or near Earth’s surface, both natural and anthropogenic; these processes can enhance the environmental mobility of metals or lead to their sequestration. Human activities have been especially effective in redistributing metals and modifying their form, including speciation. There are numerous physical, geochemical, and biologic processes that influence the behavior of metals in surficial settings. Much of the attention has focused on processes that mobilize or immobilize metals, since mobility generally facilitates exposure. A dissolved metal can be transported by fluid flow, eventually entering a water supply or being taken up within a food chain. In contrast, a metal that precipitates as a coating on mineral grains in a soil or aquifer is immobilized, and may be effectively eliminated from exposure unless remobilized by a subsequent process. In many circumstances, bioaccessibility in geomaterials is equated to the occurrence of a metal in a dissolved form. However, there are examples in which solid forms may have significant potential for exposure. Metals that are associated with small particles that are mobile may have greater potential for exposure. For example, colloids and airborne particles are mobile and both have been shown to have associated metals. Metals contained in the bottom sediments in streams, lakes, and marine settings may also enter the food chain if they are scavenged by bottom feeders or may even become remobilized upon interacting with gut juices.


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The most important processes that control the mobility of metals in the environment include dissolution/precipitation, complexation with ligands, sorption/desorption by solids, biotransformation, uptake by soil and aquatic biota, and reduction/oxidation (redox) (Fig. 3). We provide only a brief overview of these processes here, as there are a number of existing sources that offer detailed reviews relating to metals (e.g., O’Day 1999; Traina and Laperche 1999; Brown and Parks 2001;Warren and Haack 2001; Sparks 2003). Dissolution and precipitation involving metal species are subject to both thermodynamic and kinetic controls. Using the chemical analysis of a water as input, aqueous speciation programs, such as MINTEQA2 (U.S. EPA), PHREEQC (Parkhurst and Appelo 1999), and The Geochemists Workbench (http://www.rockware.com) (Bethke 2002), facilitate calculation of saturation states of the water with respect to a wide range of minerals and solids. However, there are many important environmental phases for which thermodynamic stability data are lacking, including many amorphous phases. Moreover, some of the phases controlling metal solubility in nature are solid solutions, and satisfactory solution models are available for relatively few environmentally relevant phases. Owing to a variety of possible kinetic factors, it is important to recognize that supersaturation does not guarantee that precipitation will occur. For example, inhibitors, often sorbed metals, may introduce kinetic constraints that limit precipitation (and dissolution). Plumlee et al. (2006) and Plumlee and Ziegler (2003) describe kinetic factors that influence the dissolution of asbestos and other mineral phases in human fluids. Metal solubility is also strongly coupled to complexation in the aqueous phase. Metals that have a high affinity for either organic or inorganic ligands may exhibit significantly increased solubility through formation of complexes. The high affinity of UO22+ for dissolved CO32− is an environmentally important example. In the absence of dissolved CO32−, the total solubility of UO22+ controlled by equilibrium with the mineral schoepite (UO3⋅2H2O) at neutral pH is 3.4 µM. With 1 mM total dissolved carbon dioxide in the solution the total UO22+ solubility is tenfold higher (56 µM) because of carbonate complexation. Aqueous complexation may also influence the uptake of metals by aquatic and soil organisms. For example, dissolved Cu2+ and Ni2+ are readily taken up by some aquatic phytoplankton, whereas uptake is extremely limited when these same metals are strongly complexed with organic ligands or natural organic matter. Although many studies of aquatic organisms have shown correlations between metal uptake and the fraction of the metal occurring as a free ion

Precipitation/ dissolution

Sorption/ desorption

Free metal species

Complexed metal species

Uptake/release by biota

Biotransformation/ biodegradation

Figure 3. Dominant processes that control the mobility of metals in the environment. Oxidation/reduction may be associated with any of these processes, depending on the metal.

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(i.e., the free-ion activity model)(Morel 1983; Campbell 1995), relatively little is known of the actual chemical form of the metal taken up (cf. Hudson 1998; Sunda and Huntsman 1998). In systems that contain solid phases, sorption processes, involving transfer of dissolved metals to solid surfaces, represent some of the most important controls on dissolved metal concentrations (e.g., Brown et al. 1999; Sparks 2003). Sorption can be considered to include adsorption (i.e., accumulation of ions at the surface), surface precipitation (formation of a distinct phase at the surface), and co-precipitation (incorporation of ions into a phase, commonly as it precipitates). Sorption is generally most effective at removing a dissolved metal from solution when the metal is present at relatively low concentration and the available solid surface area is high. However, sorption also depends on several solution properties, including pH, ionic strength, and the presence of complexing ligands or competing species. Properties of the solid-liquid interface, including surface charge (Hochella and White 1990) and surface site coordination (Reeder 1996; Elzinga and Reeder 2002) are also important factors. Generally, metal cations exhibit increasing adsorption as pH increases, as a response to decreasing proton charge on the surface. The change in adsorption efficiency typically occurs over a narrow pH range and is commonly referred to as the adsorption edge. The position of this edge with respect to pH may vary for different metals on the same sorbent, which generally reflects differing affinities of the metals for surface sites or different sorption mechanisms. For a given metal cation sorbing on different solid phases the pH range of the adsorption edge typically differs because of different surface charge properties and different surface sites. Adsorption of anions is usually greatest at low pH and decreases with increasing pH, also reflecting electrostatic properties of the surface. Anion species of acids may exhibit sorption maxima in their pH dependence, usually coinciding with their pKa values and demonstrating differences in sorption behavior among the more and less protonated species. Metal cation sorption may also be influenced by complexation, which may vary with pH and ligand concentration. Uranium again provides a good example. U(VI), occurring as the UO22+ aqueous species, shows a rapid increase in adsorption on ferrihydrite with increasing pH (in the pH range 3.5-5.5), which is typical behavior for a cation (Fig. 4). However, at higher pH (7.5-9.5), U(VI) adsorption decreases abruptly as a result of the formation of uranyl carbonate anion complexes having low affinity for the surface (Waite et al. 1994). Increasing the dissolved carbonate concentration causes the edge at higher pH to shift to lower pH values due to increased carbonate complexation, thereby resulting in a narrower pH range of maximum adsorption. Spectroscopic investigations have demonstrated that different types of surface complexes may form, including inner-sphere and outer-sphere types (Fig. 5). Although multiple factors may be important, inner-sphere surface complexes are generally more strongly bound to surfaces than outer-sphere complexes, and therefore may be less susceptible to desorption, or release to solution. For more information, readers are referred to the excellent reviews of sorption by Anderson and Rubin (1981), Davis and Kent (1990), Stumm (1992), Sparks (2003), and Sposito (1984, 2004). Biotransformations commonly involve reduction/oxidation (redox) processes. Because large solubility differences sometimes exist between different oxidation states of a metal, bacterially mediated reduction or oxidation may be highly effective in controlling metal concentrations in environmental solutions. Uranium again provides a useful illustration. Over the environmentally relevant pH range, U(IV) is relatively insoluble, often forming oxide phases such as uraninite, UO2. Uranium(VI) is relatively soluble and is generally considered the mobile form of uranium. Bacteria present in soil systems have been shown to reduce dissolved U(VI) to U(IV), resulting in the formation of uraninite (e.g., Lovley et al. 1991; Fredrickson et al. 2000). This process may effectively immobilize uranium in the subsurface. However, in oxidizing conditions, uraninite may become re-oxidized, resulting in its remobilization as U(VI) (e.g., Senko et al. 2002).


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Figure 4. Uranium(VI) adsorption behavior on ferrihydrite as a function of pH and at different total CO2 concentrations. The decrease in adsorption above pH 7.5-8 is attributed to formation of aqueous uranyl carbonate complexes that have low affinity for sorption. [Reprinted with permission from Waite et al., Geochimica et Cosmochimica Acta, Vol. 58, Fig. 6, p. 5470. Copyright (1994) Elsevier.]

Figure 5. Diagrams showing outer- and inner-sphere adsorption complexes and a surface precipitate. [Reproduced from Brown (1990).]

Redox processes may also occur without biological mediation, however, kinetics are commonly sluggish. In surface environments important electron donors/acceptors include organic matter and compounds containing iron, manganese, and sulfur. Many redox processes occur at surfaces of solids and are associated with sorption processes. Adsorption of metals onto surfaces can change their electronic structure and promote redox reactions that are inhibited when the metal is dissolved. For example, it has been demonstrated that reduction of Cr(VI) co-adsorbed with Fe(II) onto an Fe(III)-oxide mineral substrate is significantly faster compared to the homogenous reaction between these two metal species (Buerge and Hug 1999). The same phenomenon has been observed for the reaction between Fe(II) and uranyl (Liger et al. 1999). While the effect of sorption on the kinetics of electron transfer reactions of transition metals in aquatic systems has received some attention, this phenomenon is likely to also play a role within the human body where transition metals may be ingested or inhaled as adsorbed species on mineral dust (Schoonen et al. 2006). Further information on electron transfer reactions in the environment is provided in reviews by Bartlett and James (1993), Sparks (2003), and Schoonen and Strongin (2005).

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Because the processes described above are the dominant controls on metal mobility in the near-surface environment, they are also among the most important factors governing the environmental exposure of humans to metals. Next we consider the processes associated with metals following exposure.

Internal processes After ingestion, inhalation, or dermal contact, the fate and impact of a metal are affected by absorption, distribution, metabolism (or biotransformation), and elimination. Collectively, these processes, commonly referred to as ADME, define the field of toxicokinetics (Guidotti 2005). The behavior of metals throughout these internal processes is strongly dependent on speciation. Moreover, metal speciation typically changes during these processes. The health effect of a metal toxicant, including the mechanism of toxicity, is also dependent on speciation and constitutes the field of toxicodynamics. We do not discuss toxicodynamics, except to note a few well known examples. Moreover, the level of material that we present here is very basic, and interested readers are encouraged to consult more comprehensive reviews. A useful starting point for geoscientists is the Environmental Health and Toxicology web page of the National Library of Medicine (National Institutes of Health) (http://sis.nlm.nih.gov/enviro.html). Absorption involves transport of the metal across a physiological membrane (e.g., commonly a phospholipid bi-layer). This may occur via different mechanisms (e.g., passive, facilitated, or active transport) depending on the substance and its chemical form as well as the cell type (Dawson and Ballatori 1995; Foulkes 2000). Absorption of Cr(VI), for example, occurs via facilitated transport, following the sulfate and phosphate pathway, whereas Cr(III) transport is primarily by passive diffusion and much less efficient. Methylmercury has a significantly greater absorption efficiency than inorganic ionic mercury, which is mainly attributable to the higher lipid solubility of the methylated form. Some of the properties of metals and other substances that influence absorption are listed in Table 2. Metal speciation may be altered before absorption occurs, for example in the presence of gastric or lung fluid, or in mucus. Foulkes (2000) has emphasized that, because of their high affinity for complexing with proteins and other biological molecules in internal fluids, most non-essential heavy metals are

Table 2. Properties of a metal or substance that influence absorption at physiological membranes. Property

Effect

Concentration/dose

Fractional absorption may vary with concentration, including inversely.

Molecule size and charge

Dependent on transport mechanism (passive, facilitated, or active). For passive diffusion, neutral charge and small size favor absorption.

Competing antagonists

Competition depending on transport mechanism.

pKa of acids

Nonionized form of some acids more readily absorbed.

Lipid solubility

Lipid-soluble species more readily absorbed.

Particle size/surface area (solid)

Smaller particles solubilized more rapidly.

Phase identity (solid)

Relates to solubility and association of metal.

Solubility (solid)

More soluble solids generally allow greater absorption.

Sorption state (solid)

Metal sorbed on solid surface more readily released.

Matrix components (solid)

Other components in solid may enhance/reduce absorption.


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transported across membranes as complexes, rather than as free metals. Transformations may also include changes in metal oxidation state. An example is the partial reduction of Cr(VI) to Cr(III), which begins in the saliva and gastric fluid and influences absorption. Absorption may be influenced by a number of factors, including the fed state. For example, in rats orally ingested Cr(VI) is absorbed more readily in the fasted than in the fed state (O’Flaherty 1996). After absorption, a metal undergoes distribution and metabolism (also known as biotransformation). Some metals are distributed to all or most tissues and organs (e.g., arsenic), whereas others concentrate in specific organs, which may not be the target organs. Lead, for example concentrates in bone, yet its primary toxicity is in brain function. Both transport and storage depend on chemical form. Methylmercury, for example, readily crosses the blood-brain barrier, whereas inorganic mercury (Hg2+) does not. Metal transport in the blood commonly involves an equilibrium partitioning between plasma and proteins, with some metal species preferentially entering red blood cells. More than 90% of the methylmercury in the blood enters red blood cells and binds with hemoglobin (NRC 2000). Metals absorbed from the gastrointestinal tract do not pass directly into general systemic circulation, but first enter portal circulation where some toxicants are metabolized by the liver. Metabolism encompasses all of the biotransformations that modify the form (i.e., speciation) of a toxicant. These are typically enzymatic processes occurring with multiple steps and intermediate metabolites. Metabolites may be more or less biochemically active than the original substance. Generally metabolism serves to detoxify a toxicant, often by creating a chemical form that is more readily eliminated. For example, the methylation of arsenic in the liver facilitates its elimination and is considered a detoxification mechanism. Elimination of metals occurs primarily through urine, feces, and by exhalation. Water soluble forms are excreted readily in urine. The liver secretes some metals (e.g., lead and mercury) into bile, which are then eliminated with feces or remain in enterohepatic circulation (Guidotti 2005). Some metals are stored in various tissues, and may provide evidence of exposure over time. For example, analysis of hair serves as a method for assessing chronic exposure to mercury. The primary storage site for lead is in bone and its analysis serves as a cumulative biomarker for exposure. These ADME processes are usually unique to the particular species of metal, and clearly involve complex processes. Quantitative modeling of these processes is a major focus in the toxicokinetics field. Such physiologically based pharmacokinetic (PBPK) models have been developed for a number of metal toxicants, and are being improved as new data permit (O’Flaherty 1998). Inasmuch as quantitative modeling of environmental processes for metals is also well advanced, it may not be long before comprehensive models are developed that address both environmental and internal fate of metals. The range of environmental and internal processes that we have described is shown conceptually in Figure 6. In the following section we illustrate the ways in which speciation influences bioaccessibility and bioavailability through four examples of metals/metalloids: arsenic, chromium, lead, and mercury.

ROLE OF METAL SPECIATION: ARSENIC Arsenic has become the focus of worldwide concern following the recognition of its widespread and devastating health impact in many developing regions, most notably Bangladesh and West Bengal. The occurrence of arsenic in soils and aquatic systems, including the sediments and groundwater in the Bengal Basin region, and the geochemical processes that influence arsenic mobility have been reviewed recently by Plant et al. (2005) and Smedley and Kinniburgh (2002, 2005). Recent reviews of arsenic toxicity and human health effects

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External


total metal in system

soluble fraction

Environmental processes solid fraction

Not bioaccessible Exposure

gastrointestinal tract

liver

lungs

skin

blood and lymph circulation

Absorption

metabolism (metabolites)

bile

Internal

storage

kidney

lung

organs

feces

extracellular fluids Distribution/ Metabolism

urine

expired air

Elimination

Figure 6. Idealized diagram illustrating the range of pathways and processes affecting the environmental and physiological behavior of metals. [Adapted from the National Library of Medicine (http:// www.sis.nlm.nih.gov/enviro/toxtutor/Tox2/a11.htm).]

are given by NRC (1999, 2001), IPCS (2001), and ATSDR (2005a). Here, we provide a brief overview of the role of arsenic speciation as it relates to mobility in the environment, human exposure, bioavailability, and health effects.

Arsenic in the environment Arsenic is not an abundant element in Earth’s crust, with an average concentration in crustal rocks of ∼2 ppm (Wedepohl 1995). However, it has a strong association with sulfidebearing mineral deposits and especially with pyrite (FeS2), which is widespread. Arsenic also exhibits an association with iron oxide and hydroxide minerals, which are abundant as weathering products. Common sources of arsenic in the environment are listed in Table 3. Arsenic occurs in the 3−, 1−, 0, 3+ and 5+ oxidation states, but in nature 1−, 3+, and 5+ oxidation states dominate. Arsenide minerals and sulfides in which As(I−) is a common substituent typically exhibit low solubilities in reducing natural waters. However, oxidative weathering results in the formation of arsenite [As(III)] and arsenate [As(V)] species; these occur as oxyanions or neutral species and may be quite soluble depending on pH and other solution properties. In solution, redox potential (Eh) and pH are the dominant controls on As speciation. An Eh-pH stability diagram is shown in Figure 7. Coexistence of As(III) and As(V) species in natural waters has been interpreted to indicate that electron transfer occurs, if not redox equilibrium. However, rates of electron transfer may be slow and strongly influenced


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Table 3. Common sources of arsenic in the environment. • Volcanic emissions and hot springs • Dissolved in groundwater from interaction with rock (mobilization from igneous and sedimentary sources, oxidative dissolution of arsenic-bearing sulfide minerals) • Mining waste, pH-mediated mine effluents, and tailings ponds • Arsenic-containing pesticides (sodium arsenite or lead arsenate) • Organic arsenic compounds as herbicides: monosodium methanoarsonate (MSMA), disodium methanoarsonate (DSMA), arsenic acid, and dimethylarsenic acid • Waste from industrial metal smelting processes • Leaching of wood preservatives: chromated copper arsenate (CCA) and ammoniacal copper arsenate • Combustion of fossil fuels in electrical power plants • Arsenic dusts and gases released during cement manufacture • Animal waste management from feed additives in poultry (roxarsone to control coccidiosis and promote growth) • Arsenic trioxide waste from glass manufacturing process (pre-1970)

by bacterial activity (e.g., Cullen and Reimer 1989; Smedley and Kinniburgh 2005). The acid dissociation constants (Ka), expressed here as pKa values, for H3AsO4 are 2.22, 6.98, and 11.53. Below pH ∼7 the dominant As(V) species is H2AsO4− and above pH ∼7 HAsO42− is dominant, except at very low or high pH (Fig. 8). The pKa values for H3AsO3 are 9.23, 12.13, and 13.4, so that H3AsO30 is the dominant dissolved inorganic As(III) species in most natural waters. In view of its affinity for sulfur, As(III) may form thio-complexes (cf. Wilkin et al. 2003). Methylation may also occur in the environment as a result of microbial processes, resulting in formation of several As(III) and As(V) methyl species, including monomethylarsonous acid [MMA(III)], dimethylarsinous acid [DMA(III)], monomethylarsinous acid [MMA(V)], and dimethylarsinic acid [DMA(V)] (Fig. 9). Anderson and Bruland (1991) reported the formation of dimethylarsenate [DMA(V)], (CH3)2AsO2−, on a seasonal basis to become the dominant dissolved As species in a freshwater reservoir, followed by breakdown to arsenate. Bright et al. (1996) also reported methylated arsenic species in lake sediment pore waters, with formation presumed to be associated with bacterial activity. Other organic species include arsenobetaine and arsenocholine, which are forms commonly present in food. In the majority of natural waters, however, inorganic arsenic forms are most abundant (Smedley and Kinniburgh 2002). Human exposure to arsenic is primarily through ingestion of water and food, although airborne particulates containing arsenic may be locally important. For this example, we focus on exposure through drinking water and to a lesser extent on ingestion of arsenic-containing solids. Consequently, solubility of arsenic is the main geochemical factor influencing exposure in this circumstance. Additionally, arsenic associated with particles that are mobile could also contribute to exposure. As noted earlier, numerous geochemical processes control metal solubility and mobility, including mineral precipitation/dissolution, sorption, and biotransformation. Metal speciation is a critical aspect of all these processes. Arsenic sorption. Because dissolved arsenic concentrations are typically very low, even in contaminated systems, sorption processes can be highly effective in limiting As concentration and thereby controlling mobility and bioaccessibility. A number of studies have demonstrated that As(III) and As(V) species may sorb strongly with oxide and hydroxide minerals, especially iron, aluminum, and manganese oxyhydroxides. Sorption is strongly dependent on pH and

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1 H3 AsO4 -

H2AsO4

Eh (volts)

.5

HA sO

Figure 7. Eh-pH predominance diagram for arsenic in the presence of sulfur at 25 °C (Astot = 10−6 M; Stot = 10−4 M).

2-

HAsO4

2 (aq

)

0

3-

AsO4 Orpiment

-

AsO2

–.5

As O

25°C

0

2 OH 2 -

As

2

4

6

8

10

12

14

pH

a

b H2AsO3-

100 H3AsO3

100

80

% As(V) species

% As(III) species

HAsO42-

H2AsO4-

80

60

40

HAsO32-

20

60

40

AsO43-

20 H3AsO4

3-

AsO3 0

0 3

4

5

6

7

8

pH

9

10

11

12

3

4

5

6

7

8

9

10

11

12

pH

Figure 8. Aqueous speciation diagrams for (a) As(III) and (b) As(V) systems at 25 °C (Astot = 10 ppb, and 1 mM NaCl).

other solution properties. Arsenite and arsenate exhibit very different sorption behaviors as a function of pH, and in some systems sorption is sensitive to ionic strength. As an example, Arai et al. (2001) compared As(III) and As(V) sorption on γ-alumina over the pH range 3-10 (Fig. 10). Arsenate sorption increases with decreasing pH, over the range from pH 9.5 (near the point of zero charge, PZC) to 4.5. This behavior is typical for sorption of anion species (e.g., Hingston 1981). Arsenate sorption is not observed to be sensitive to ionic strength; this has been interpreted to be suggestive of formation of inner-sphere surface complexes, which was confirmed using EXAFS spectroscopy (Arai et al. 2001). Arsenite exhibits less overall sorption than arsenate on γ-alumina, with a broad maximum at pH ∼8.5 (see also Goldberg 2002). A decrease in arsenite sorption is observed above pH 9 and likely reflects the change in dominant aqueous speciation to an anion, H2AsO31−. Over

Role of Metal Speciation in Bioaccessibility & Bioavailability O xx

As HO

OH OH

Arsenious Acid (arsenite)

As HO

OH OH

Arsenic Acid (arsenate) O

As H3C

OH OH

Monomethylarsonous acid MMA(III)

As H3C

OH OH

Monomethylarsonic acid MMA(V) O

As H3C

CH3 OH

Dimethylarsinous acid DMA(III)

As H3C

OH

CH3

Dimethylarsinic acid DMA(V)

Figure 9. Common inorganic and methylated As(III) and As(V) species. [Adapted from O’Day (2006).]

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the pH range 5-9, arsenite sorption is decreased at higher ionic strength, which is suggestive of outer-sphere surface complexation. In situ EXAFS spectroscopy showed that inner-sphere As(III) complexes dominate at pH < 5.5 and a mixture of inner- and outer-sphere complexes exists at pH > 5.5 (Arai et al. 2001). Several studies have shown that arsenate and phosphate compete for sorption sites, owing to similar chemical behaviors (e.g., Jain and Loeppert 2000). Consequently arsenate sorption may be significantly depressed where phosphate is elevated. There are numerous studies of As(III) and As(V) sorption on other mineral and oxide surfaces that readers should consult for further insight (e.g., Fuller et al. 1993; Waychunas et al. 1993; Goldberg 2002; Smedley and Kinniburgh 2002; Stollenwerk 2003). Arsenic redox behavior. Much of the arsenic present in surface environments has been derived from the oxidative weathering of reduced species in sulfide mineral deposits and primary rocks (Smedley and Kinniburgh 2002). Because of the relatively low solubility of reduced arsenic, much of the interest in redox behavior, from the perspective of health impact,

Figure 10. The pH and ionic strength dependence of As(III) and As(V) sorption on γ-alumina. [Reprinted with permission from Arai et al., Journal of Colloid and Interface Science, Vol. 253, Fig. 1, p. 83. Copyright (2001) Elsevier.]

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has focused on transformations between As(III) and As(V) species. Both oxidation states are common in soils and near-surface waters, and it is not uncommon to have both As(III) and As(V) species co-existing in solutions and in solids (Hering and Kneebone 2002). The EhpH diagram shows that the predominance fields for species of both oxidation states coincide with Eh-pH conditions typical of many environmentally important settings (Fig. 7), so that redox transformations can be expected. However, in most cases As(III)/As(V) redox behavior is controlled by interaction with minerals, microbes, or organic matter that serve as electron donors or accepters. As is often the case in systems containing multiple redox couples, the As(III)/As(V) couple is commonly not in equilibrium with other redox couples (Spliethoff et al. 1995; Hering and Kneebone 2002). This reflects widely differing kinetics of electron transfer and strong dependence on mechanism. Hence As(III) species may persist in oxidizing systems and As(V) may persist in reducing conditions (cf. Inskeep et al. 2002). Important oxidants for As(III) include Mn(IV) oxides (e.g., birnessite), titanium oxides, H2O2, and possibly dissolved ferric species (Foster et al. 1998; Manning et al. 2002; Voegelin and Hug 2003). Photochemical oxidation may also be important (Inskeep et al. 2002). Dissolved oxygen has been shown not to be an effective oxidant, except at high pH (e.g., Manning and Goldberg 1997). Dissolved sulfide has been shown to reduce As(V), with formation of intermediate sulfide complexes (Rochette et al. 2000). Many of the important redox processes occur at mineral surfaces and are associated with sorption. As noted above, arsenate is often associated with iron and aluminum hydroxides, and is typically more strongly sorbed by these phases than arsenite at circum-neutral and acidic pH conditions. An important mechanism of release of sorbed As(V) involves its reduction to As(III). This may occur either by arsenate reduction at the surface with subsequent As(III) release to solution, or by reductive dissolution of a ferric hydroxide sorbent followed by reduction of arsenate to arsenite as shown in Figure 11 (Inskeep et al. 2002). Reductive dissolution of iron (hydr)oxides (with sorbed arsenate) is thought to be a factor causing the elevated dissolved arsenic concentrations in groundwater in Bangladesh (e.g., Smedley and Kinniburgh 2002). Microbial activity has been shown to cause both reduction and oxidation of arsenic species. We do not describe these reactions here, but refer readers to the overview of Inskeep et al. (2002). Arsenic precipitation/dissolution. Precipitation and dissolution may also provide important constraints on As solubility. In many aquatic systems, however, arsenic concentrations

Figure 11. Schematic diagram showing reductive release mechanisms for As(III). [From Inskeep et al. (2002), Environmental Chemistry of Arsenic. p. 191, Fig. 4. Reproduced by permission of Routledge/ Taylor & Francis Group LLC. Copyright (2002).]


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remain very low so that supersaturation with respect to arsenic oxides/hydroxides or to arsenite or arsenate salts is uncommon. However, high concentrations of cations may allow supersaturation and precipitation. Some remediation strategies rely on precipitation to remove arsenic from solution, such as wastewater. Additives such as lime, fly ash, Portland cement, and ferrihydrite have been used to induced formation of arsenic phases (e.g., Moon et al. 2004). Calcium arsenite and arsenate phases have received some attention as possible precipitates for immobilizing arsenic (e.g., Bothe and Brown 1999). However, the effectiveness of precipitation as a remediation strategy depends on the solubility of the solid. Overviews of important arsenic-containing phases and minerals are given by Cullen and Reimer (1989), Nordstrom and Archer (2003), and O’Day (2006). Foster (2003) describes arsenic speciation in a number of solid phases, including that for arsenic bound with oxides and hydroxides. Arsenic mineral solubilities vary widely according to arsenic oxidation state, composition, crystal structure. The total dissolved arsenic concentration and the concentrations of individual arsenic species in equilibrium with different arsenic-containing phases depend on the overall solution composition, including pH, ionic strength, T, and P. Aqueous speciation programs facilitate such calculations and provide an equilibrium distribution of aqueous species. Nordstrom and Archer (2003) have critically evaluated the thermochemical data available for selected arsenic phases and aqueous species. In general, As(III) and As(V) oxides are moderately soluble in most aqueous solutions. As2O3 occurs as arsenolite and claudetite, whereas As2O5 is not known as a mineral. Arsenic sulfides, such as realgar and orpiment, tend to be relatively insoluble in most reduced solutions, as are metal arsenides and arsenic sulfosalts. Metal arsenates and arsenites exhibit a range of solubilities depending on the metal involved and other factors (cf. Sadiq 1997). As described below, solubility is important in controlling the bioavailability of arsenic-containing solids following ingestion or inhalation. Arsenic coprecipitation. Arsenic may also coprecipitate with mineral phases that are forming, thereby providing an effective mechanism for removing arsenic from solution. For example, coprecipitation with ferrihydrite has been shown to significantly reduce dissolved arsenate (e.g., Fuller et al. 1993; Richmond et al. 2004). The similarity of AsO43− to other tetrahedral oxoanions, such as PO43− and SO42−, should allow As(V) substitution (e.g., Myneni et al. 1997; Foster 2003). In addition As commonly substitutes for S in pyrite and other sulfides.

Arsenic in the body Arsenic absorption. Once ingested, dissolved As(III) and As(V) are both readily absorbed in the human gastrointestinal tract. Studies that systematically compare the relative absorption efficiencies of dissolved As(III) and As(V) in humans from oral exposure are lacking. However, an NRC review cited studies of humans and animals indicating 80-90% absorption for dissolved As(III) and As(V) doses (NRC 1999). MMA(V) and DMA(V) are also readily absorbed in the gastrointestinal tract of humans (ATSDR 2005a), as is arsenobetaine from ingestion of fish (IPCS 2001). Arsenic present in or associated with solid forms generally shows lower absorption efficiency than dissolved arsenite or arsenate, especially for phases exhibiting low solubility. There are few systematic studies of gastrointestinal absorption (oral bioavailability) in humans for different As compounds. Studies in animal models have been summarized by ATSDR (2005a) and IPCS (2001). Although significant differences exist between different animals, solubility of the arsenic source is the most significant factor influencing absorption. For example, gastrointestinal absorption in rodents is low for relatively insoluble GaAs ( As(V) >> MMA(V) ≅ DMA(V) (and other organic forms). Ingested MMA(V) and DMA(V) are metabolized less and eliminated more rapidly and to a greater degree than inorganic arsenic (IPCS 2001; ATSDR 2005a). Arsenic present in fish, primarily arsenobetaine, apparently undergoes little metabolism and is readily eliminated (Le 2002). At high doses, studies have shown that more arsenic is retained after ingestion of As(III) than As(V) (NRC 1999). A significant factor contributing to the greater toxicity of As(III) is its greater solubility in lipids and its ability to cross cell membranes more readily than As(V) species (Schoolmeester and White 1980), in part due to its neutral charge at physiological pH. Studies using laboratory animals have generally shown lower LD50 concentrations for As(III) ingestion compared to As(V) ingestion (IPCS 2001), which supports the greater toxicity of the As(III) form. As noted above, there is recent evidence suggesting that the toxicity of the metabolite MMA(III) may be more toxic that inorganic As(III). However, the ATSDR (2005a) has emphasized that studies based on laboratory animals do not provide good quantitative models for human toxicity. Finally, many related factors are known or thought to modify arsenic toxicity. For example, diet has been shown to influence arsenic toxicity (Peraza et al. 1998) and more specifically the degree of methylation (e.g., Steinmaus et al. 2005), which may be one factor that explains variability of health effects among individuals or groups. Another aspect of interest is the observation that co-exposure to selenium may reduce the toxicity of arsenic (e.g., Levander 1977). The mechanism for this effect is under investigation, but may involve the formation of an arsenic-selenium complex either with reduced toxicity or more rapid elimination. Recent studies by Gailer and co-workers (2000, 2002) have identified the formation of a seleno-bis (Sglutathionyl) arsinium complex, [(GS)2AsSe]−, in rabbits injected with arsenite and selenate, with rapid excretion to bile.

ROLE OF METAL SPECIATION: CHROMIUM Chromium is one of the more abundant heavy metals, with an average crustal concentration of 126 ppm (Wedepohl 1995). There is a long history of mining and processing of chromium, fueled by numerous industrial applications. Chromium is an important component in steel and other alloys, paints, magnetic recording tape, electroplating, wood preservative, and leather tanning, and serves as an anticorrosive agent in water-cooling systems. Other sources of chromium in the environment are listed in Table 4. Its widespread use has been accompanied by releases to the environment that pose persistent health hazards. Chromium is among the metals included in the ATSDR/EPA National Priority List of hazardous substances, and is present at a majority of sites on the CERCLA National Priority List (http://www.atsdr.cdc.gov/cercla). The environmental behavior and toxicology of chromium have been studied extensively. Readers are directed to the reviews of the environmental behavior of chromium by Rai et al. (1989), Fendorf (1995), and Kimbrough et al. (1999). Summaries of the environmental toxicology of chromium are given in ATSDR (2000), IPCS (1988), EPA (1998a,b), and O’Flaherty et al. (2001).

Chromium in the environment Chromium may exist in oxidation states ranging from −II to VI. In the near-surface environment, the III and VI oxidation states are most important. It is well known that Cr(III)

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Reeder, Schoonen, Lanziro i Table 4. Some common sources of chromium in the environment. • Occupational exposure from production of chromate, stainless-steel, chrome plating • Air emissions and water effluents from ferrochrome production, ore refining, tanning industries, chemical manufacturing industries (e.g., dyes for paints, rubber and plastic products), metal-finishing industries (e.g., chrome plating), manufacturing of pharmaceuticals, wood, stone, clay and glass products, electrical and aircraft manufacturing, steam and air conditioning supply services, cement production • Air emissions from incineration of refuse and sewage sludge • Combustion of oil and coal • Oxidation and leaching from stainless steel into a water-soluble form • Motor vehicle exhaust and emissions from automobile brake linings and catalytic converters • Tobacco smoke

is an essential micronutrient, playing a role in glucose metabolism. In contrast, Cr(VI) is considered a known carcinogen (by inhalation route) and is associated with both acute and chronic health effects (by inhalation and ingestion), as well as being a contact allergen (EPA 1998a; ATSDR 2000). Cr(IV) and Cr(V) are thought to play a role in the mechanism of toxicity and have been associated with production of reactive oxygen species and carbonbased radicals that may interact with DNA (cf. EPA 1998a; Gaggelli et al. 2002). In minerals, water, and most soils, chromium is dominantly coordinated with oxygen. Cr(III) shows a strong preference for octahedral coordination. The Cr3+ aqua ion undergoes hydrolysis, yielding Cr(OH)2+, Cr(OH)2+, Cr(OH)30, and Cr(OH)4− species depending on pH (Fig. 13). Cr(III) may also form aqueous complexes with organic and inorganic ligands. Cr(VI) occurs almost exclusively in tetrahedral coordination with oxygen, as the chromate anion (CrO42−). The pKa2 value for chromic acid, H2CrO4, is ~6.5, so that HCrO4− dominates in solutions below pH 6.5, and CrO42− dominates above (Fig. 13). Cr(VI) also occurs as dichromate, Cr2O72−, but this species only become important at millimolar Cr concentrations and greater.

1 -

.5

3+

Cr

2-

2+

CrO4 CrOH

0

Figure 13. Eh-pH predominance diagram for aqueous chromium at 25 oC (Crtot = 1 µM).

3-

Cr(OH) 3 (am)

CrO4

-

Cr(OH)4

Eh (volts)

HCrO4

–.5 25°C

0

2

4

6

8

pH

10

12

14


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Interconversion between Cr(III) and Cr(VI) depends on oxidation potential and pH (Fig. 13). Inasmuch as chromium is usually present at very low concentration in soils, surface sediments, and aquatic systems, other redox couples, commonly involving iron or organic matter, are usually important in controlling chromium oxidation state. Chromium occurring in natural minerals is dominantly as Cr(III), as in the important ore mineral chromite (FeCr2O4). The occurrence of Cr(VI) species and compounds is usually a result of anthropogenic activities or oxidative weathering. Solubility and dissolution/precipitation. There are significant differences in the behavior of Cr(III) and Cr(VI) species under most surface conditions. Cr(III) compounds and minerals are largely insoluble except at low and very high pH. Over the pH range 6-10.5 precipitation of crystalline or amorphous Cr(OH)3 or (Fe,Cr)(OH)3 effectively limits dissolved Cr(III) concentrations to values below the current EPA drinking water MCL of 0.1 mg/L (Rai et al. 1987; Sass and Rai 1987), thereby limiting the oral exposure of this Cr species through water. In contrast, Cr(VI) is highly soluble over the pH range of most natural waters. Chromate salts of sodium, potassium, magnesium, and calcium are highly soluble and rarely limit environmental Cr(VI) concentrations. Only PbCrO4 (crocoite) is relatively insoluble. This striking difference in solubility between Cr(III) and Cr(VI) species has the unfortunate consequence that the essential micronutrient, Cr(III), is largely absent as a dissolved species in water while the toxic form, Cr(VI), is soluble and therefore potentially mobile. This solubility difference, however, also presents a potential remediation strategy based on conversion of soluble Cr(VI) to insoluble Cr(III). Chromium reduction/oxidation. Owing to the significant difference in Cr(III) and Cr(VI) mobilities in the near-surface environment, redox processes are important for influencing chromium bioaccessibility and exposure to organisms. Although pH dependent, the Cr(III/ VI) redox potential is relatively high, and CrO42− is a strong oxidant. Ferrous iron, organic matter, and sulfides have been shown to reduce Cr(VI) to Cr(III) readily even in the presence of dissolved oxygen (James and Bartlett 1983a,b; Rai et al. 1989; Fendorf 1995; Patterson et al. 1997). Reduction by Fe(II) has been shown to result in the formation of (FeIII,CrIII)(OH)3, which is insoluble and stable (Eary and Rai 1988). Cr(VI) reduction may be facilitated at surfaces of minerals containing Fe(II). For example, magnetite and ferrous biotite surfaces have been shown to reduce Cr(VI) to Cr(III) (Ilton and Veblen 1994; Peterson et al. 1996; Peterson et al. 1997). Fendorf et al. (2000) also demonstrated that Cr(VI) may be reduced by bacterial processes even in aerobic conditions. Rai et al. (1989) and Fendorf (1995) report several studies in which Cr(III) was rapidly oxidized to Cr(VI) by manganese oxides, which are common constituents in soils. The effectiveness of this process, however, may be limited by formation of a hydrous Cr(OH)3 or CrOOH precipitate at the MnO2 surface (Fendorf et al. 1992; Charlet and Manceau 1993). Chromium sorption. Owing to the high mobility of Cr(VI) in environmental systems much attention has been focused on the sorption behavior of this Cr species and its effectiveness in reducing mobility. As anion species (CrO42− and HCrO4−), Cr(VI) exhibits greatest sorption at low pH (Fig. 14). Above pH 7 sorption may not be effective in removing Cr(VI) from solution (Zachara et al. 1987, 1989, 2004). In view of their nearly ubiquitous occurrence, iron oxides and hydroxides are likely to be some of the most effective sorbents. Chromate sorption may be strongly diminished as a result of competition for available surface sites by other anion species, such as carbonate, phosphate, and sulfate (e.g., Zachara et al. 1987).

Chromium in the body Chromium exposure. The previous discussion focused primarily on environmental processes that control the behavior of different chromium species in soils, sediments, and aquatic systems. These influence human exposure mainly through ingestion of water, but may

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Figure 14. The pH dependence of chromate adsorption on different mineral and solid sorbents. [Reprinted with permission from Rai D et al., The Science of the Total Environment, Vol. 86, Fig. 4, p. 21. Copyright (1989) Elsevier.]

also relate to accidental ingestion of soil or inhalation of mineral dusts that contain chromium. The largest source of chromium intake for the general population is from food, where it occurs in the Cr(III) state (WHO 2003). Inhalation of Cr-containing airborne particles from industrial emissions, both Cr(III) and Cr(VI), may be locally significant, but most instances are associated with occupational exposures. Exposure may also occur through dermal contact. Here, we restrict the following discussion to illustrate examples where chromium speciation is relevant to health effects. Much of this information has been taken from the comprehensive reviews provided by IPCS (1988), EPA (1998a,b), ATSDR (2000), and WHO (2003). Readers are also referred to O’Flaherty (1996) and O’Flaherty et al. (2001). Absorption and distribution. The efficiency of chromium absorption in the gastrointestinal tract is relatively low, and contrasts with the much greater absorption of arsenic noted in the previous section. Cr(VI) exhibits greater absorption than Cr(III), but both depend on the chemical form and other factors, including food and nutritional status. After oral exposure, most studies have shown absorption of soluble Cr(III) of ~0.5%, whereas absorption of soluble Cr(VI) compounds was 2-7% (ATSDR 2000; EPA 1998a,b), with a similar absorption efficiency for dissolved Cr(VI). However, Cr(VI) is partially reduced to Cr(III) in gastric fluid (and elsewhere in the body) (De Flora et al. 1987), which decreases the total absorption owing to the lower absorption efficiency of Cr(III). Essentially no absorption was observed following oral exposure of Cr(III) as insoluble Cr2O3 (Finley et al. 1996). Absorption of dietary Cr(III) is approximately 0.5-2% (EPA 1998b), and absorption of Cr(III) in the form Cr(III) picolinate (the form commonly used in vitamins) is as much as 3% (Gargas et al. 1994). Studies using lab animals generally support the greater absorption efficiency of Cr(VI) over Cr(III). Cr(III) absorption, however, may be enhanced when Cr(III) is complexed by organic ligands (e.g., oxalate) (ATSDR 2000) or when present in biologically active complexes (Mertz and Roginski 1971; O’Flaherty 1996). Studies of inhalation exposure also show greater absorption of Cr(VI) than Cr(III) in the lungs. This has been confirmed by studies using lab animals, which have also indicated that the efficiency of chromium absorption in the lungs may be significantly greater than in the GI tract. Studies using a rat model (summarized in ATSDR 2000) showed 53-85% absorption from Cr(VI) particles (including clearance to the pharynx and into the GI tract) and 5-30%


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absorption from Cr(III) particles. Lab animal studies also show that absorption depends strongly on the solubility of the inhaled compound, with greater absorption for more soluble phases (Bragt and van Dura 1983). This supports the view that dissolution of a solid to a soluble form is normally required for absorption across a physiological membrane. The greater absorption of Cr(VI) compared to Cr(III) is largely due to the facilitated transport of chromate across cell membranes, following the same anion channel pathway as sulfate and phosphate (Wiegand et al. 1985). Absorption of Cr(III) is limited, occurring only by passive diffusion and/or phagocytotsis, which are much less effective. In the blood, Cr(VI) is able to cross the membrane of red blood cells, where it is rapidly reduced to Cr(III) and interacts with proteins. In contrast, Cr(III) is largely restricted to the plasma (WHO 2003). The greater tendency of Cr(VI) to cross cell membranes is also reflected in chromium distribution following exposure. One study using a mouse model showed that Cr was detected only in the liver following one year of exposure to Cr(III) chloride (a soluble form). In contrast, Cr was detected in all organs for mice exposed to soluble Cr(VI) for the same period (ATSDR 2000). O’Flaherty (1996) has suggested that because of its low absorption efficiency the factors that influence bioaccessibility of any particular environmental chromium source are likely to be the single most important determinant of toxicity. Metabolism and elimination. The essential role of Cr(III) is associated with a biologically active Cr(III) complex that is involved in glucose metabolism (Anderson 1986). Cr(VI) is not stable in the body, and a variety of electron donors, such as ascorbate and glutathione, cause reduction to Cr(III) species (De Flora et al. 1987). This proceeds throughout the body, including in saliva, gastric and lung fluids, blood, and in major organs. This reduction of Cr(VI) can be considered a defense or detoxification mechanism. However, Cr(V) and Cr(IV) intermediates are formed during reduction; these have been associated with formation of reactive oxygen species, and may be involved in the mechanism of toxicity (Gaggelli et al. 2002; Levina et al. 2003). Absorbed chromium is mostly excreted as Cr(III) complexes through the urine within a period of several hours to several days (ATSDR 2000). The absence of Cr(VI) in urine, even after exposure to Cr(VI), indicates that reduction is complete within this time frame. Some chromium is retained in tissue and bone for periods on the order of months or longer. The large chromium fraction that is not absorbed following oral exposure is eliminated in the feces.

ROLE OF METAL SPECIATION: LEAD Lead in the environment Lead poisoning is one of the most common and serious environmental issues in industrialized nations, particularly with regard to its effect on the cognitive development of young children. Although lead occurs in the environment naturally, the vast majority of the instances of elevated lead levels in the environment that are of concern for human health are the result of human activity. Anthropogenic sources of lead include the mining, smelting, and refining of lead ore, emissions from coal and oil combustion, emissions from combustion of leaded gasoline, lead-based paints and solders, lead arsenate pesticides, and waste incineration. Much of the review of lead in the environment provided here is taken from the U. S. Department of Health and Human Service’s Agency for Toxic Substances and Disease Registry on the effects of lead in the environment and its toxicology (ATSDR 2005b). In addition to the ATSDR overview, excellent reviews are also provided by Baxter and Frech (1995) and Ryan et al. (2004). Here, we provide a brief overview of the role of lead speciation as it relates to mobility in the environment and human exposure, bioavailability, and health effects.

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Lead as a mineral is extremely rare (Wedepohl 1995). The most common form of lead in the Earth’s crust is galena (PbS), but other important ore minerals include cerussite (PbCO3), anglesite (PbSO4), and minium (Pb3O4). Lead minerals are commonly found in association with zinc, copper, and iron sulfides and are commonly associated with gold and silver ores. Lead can also occur as a trace element in coal, oil, and wood. The majority of the lead used in industry today comes from recycling. Lead occurs in both the 2+ and 4+ oxidation states, but in nature the 2+ oxidation state dominates. Lead species in natural waters. In most ground and surface waters, the solubility of lead compounds is generally low. The solubilities are dependent on water pH, the presence of coexisting ionic species and ligands, water salinity, and the organic matter content. Solubility tends to be highest in soft, acidic water. Eh-pH diagrams for the system Pb-CO2-H2O are shown in Figures 15a and 15b for 5 ppb and 207 ppb Pbtot, respectively a (the EPA action level for Pb in drinking water is 15 ppb; 207 ppb = 1 µM). Lead can exist as the Pb2+ ionic species at pH < 7.5 in fresh water, but readily complexes with dissolved carbonate at pH > 5.4 and forms lead carbonates, PbCO3 and Pb2(OH)2CO3, limiting its solubility (Long and Angino 1977). In natural waters lead will readily precipitate in the form of various lead hydroxides, carbonates, sulfates and phosphates, so that the amount of lead remaining dissolved in solution is typically low (Mundell et al. 1989). In highly oxidizing waters, such as may be found in municipal water b systems, lead may precipitate as plattnerite (PbO2). In surface waters most of the lead that is in suspension occurs as colloidal and undissolved particles. In seawater lead carbonate complexes are also prevalent, but lead chloride complexes and surface complexes with iron and manganese oxides can also occur (Long and Angino 1977; Elbaz-Poulichet et al. 1984). Lead species in the atmosphere. The largest influx of lead to the environment has historically come from emissions to the atmosphere due to automotive and industrial sources. Lead particles are emitted from smelters primarily in the form of elemental lead and lead-sulfur compounds, PbSO4,

Figure 15. Eh-pH predominance diagrams for lead in the system Pb-CO2-H2O, 25 oC. (a) Pbtot = 5 ppb (0.024 µM), Ctot = 18 ppm (1.5 mM); (b) Pbtot = 207 ppb (1 µM), Ctot = 18 ppm.


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PbO·PbSO4, and PbS (Corrin and Natusch 1977; EPA 1986; Spear et al. 1998). Oil and coal combustion has also been found to release PbCl2 and PbO (Nerin et al. 1999). Prior to 1990 nearly 90% of all lead emissions were the direct result of combustion of leaded gasoline. In 1984 the U.S. Environmental Protection Agency mandated the gradual phase-out of the use of lead alkyls in gasoline and banned their use entirely in 1996 (EPA 1996). Today, industrial emissions from metal smelting operations and battery manufacturing are the largest sources of lead to the atmosphere. The vast majority of the lead emissions from leaded gasoline are in the form of inorganic particulates such as lead bromochloride. These halogenated lead compounds are formed from the combustion of gasoline as the tetra-alkyl lead additives react with halogenated lead scavenger compounds (EPA 1985). Less than 10% of the lead emitted from combustion of leaded gasoline is emitted as organolead compounds such as vapor phase lead alkyls. However, tetra-alkyl lead vapors are strongly photoreactive, and quickly react to form trialkyl and dialkyl lead compounds, and eventually inorganic lead oxides (Eisenreich et al. 1981). In direct sunlight the half-life of tetraethyl lead vapors is approximately 2 hours (DeJonghe and Adams 1986). These lead particles are then deposited to either land or water through both wet and dry deposition. Their fine particle size and solubility in acidic solutions make these particles a particularly potent source of bioaccessible lead. Lead species in soils. Once deposited to soils, the speciation of lead can take several different forms depending on the soil type. Key factors controlling its speciation in soils include soil pH, the organic matter content, and the soil cation exchange capacity (NSF 1977; Reddy et al. 1995). All of these factors, along with the mineral form, particle size, and association with other mineral phases, affect speciation of lead in soils and its potential for solubilization, which would influence its bioaccessibility and bioavailability (Fig. 16; Ruby et al. 1999).

Figure 16. Properties of lead species and compounds and their effects on lead bioaccessibility and bioavailability in soils. [Reprinted with permission from Ruby et al. (1999). Copyright (1999) American Chemical Society.]

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Sorption processes play a very important role in lead speciation in soils. In particular, iron(III) and manganese(III/IV) oxides and hydroxides, which are ubiquitous in most soils, are very effective sorbents for lead (O’Reilly and Hochella 2003). A number of studies (e.g., O’Reilly and Hochella 2003; Villalobos et al. 2005) have demonstrated that Mn-oxides in particular are very efficient sorbents of Pb. Figure 17 shows the measured lead sorption capacities for a variety of synthetic Fe- and Mn-oxides. Synthetic birnessite, in particular, can sorb large amounts of lead from solution, almost an order of magnitude more than Fe-oxides. EXAFS studies have indicated that Pb2+ forms inner-sphere complexes within the birnessite inter-layer (Morin et al. 1999; Matocha et al. 2001; Manceau et al. 2002). Soil organic matter can also play a vital role in determining lead speciation in soils. Soil humic and fulvic substances, for example, have been shown to tightly bind lead (Pinheiro et al. 1999; Christl et al. 2005; Newton et al. 2006). Although humic substances contain several major functional groups (primarily carboxylic groups ~80% and phenolic groups ~20%), it is generally thought that Pb exhibits a strong affinity for the carboxylic groups (COOH) of the humic substances (Stevenson 1994; Xia et al. 1997). X-ray absorption studies of Pb-humate systems indicate that Pb can form inner-sphere complexes with the humic component and that the structure of the binding site does not change appreciably as a function of pH (Xia et al. 1997). For example, Morin et al. (1999) examined smelter-contaminated soils from EvinMalmaison, France, analyzing both tilled and wooded soil samples. They showed that although the source of Pb contamination was similar, differences in the chemical forms of Pb existed depending on soil type. In the wooded soil ~40% of the Pb was found to exist as outer-sphere complexes, ~10% as Pb2+ adsorbed on hydrous Fe oxides or oxyhydroxides, and ~50% as Pb2+ inner-sphere complexes bound to organic matter. By contrast, in the tilled soils Pb was present almost entirely as inner-sphere complexes adsorbed on hydrous Fe and Mn oxides. The higher amount of organo-Pb2+ and exchangeable outer-sphere Pb2+ complexes in the wooded soil was attributed to their higher organic matter content (6.4 wt% vs. 1.5 wt% TOC in the tilled soils).

Figure 17. Average total lead sorption by various Mn and Fe-oxide minerals. Studies were conducted using NaNO3 (white bars) and KNO3 (gray bars) background electrolyte. [Reprinted with permission from O’Reilly and Hochella, Geochimica et Cosmochimica Acta, Vol. 67, Fig. 6, p. 4479. Copyright (2003) Elsevier.]


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However, the binding of lead with humic components can be strongly affected by the presence of other major ions such as Ca2+, Mg2+, and Al3+. Mota et al. (1996) demonstrated that by increasing Al3+ concentration, it was possible to also increase the concentration of free lead, showing that aluminum competes with lead for the binding sites of humic substances. This can be an important effect in soils containing gibbsite, where Al3+ concentrations may be high. Pinheiro et al. (1999) demonstrated a similar competitive binding of Pb2+ and Ca2+ to fulvic acid. Whereas solid organics can be quite efficient in diminishing lead mobility, aqueous organics, by contrast, can act to enhance lead mobility. Organic acids produced by plants and soil microorganisms can be particularly effective in chelating lead and enhancing mobility in soils. Soil phosphate content also plays a role in controlling lead speciation. A number of studies of Pb-contaminated soils have shown that precipitation of Pb2+ as lead phosphate is effective in sequestering lead in soils ( Davis et al. 1993; Cotter-Howells et al. 1994; Ruby et al. 1994; Juillot 1998; Traina and Laperche 1999; Morin et al. 2001). This is particularly important in evaluating lead bioaccessibility, inasmuch as members of the pyromorphite group [(Pb,Ca)5 (PO4)3(OH,Cl,F)] are among the most stable and least soluble lead minerals under surface environmental conditions (Nriagu 1973, 1974, 1984) and may form rapidly when adequate phosphate is present (Ruby et al. 1994). Their formation effectively limits the mobility of Pb in some soils. Cotter-Howells (1996) reported on the presence of lead phosphate in urban and roadside soils, demonstrating that these phases formed in the soil as a weathering product of Pb-bearing grains. The low solubility of pyromorphite has also advanced the use of phosphate amendments to lead-contaminated soils to promote pyromorphite formation. There is also some indication that soil organisms may be able to biogenically precipitate pyromorphite in contaminated environments. Jackson et al. (2005) demonstrated that soil nematodes placed in a KCl-NaCl medium along with Pb(NO3)2 immobilize the Pb through pyromorphite precipitation. Given the high density and turnover of roundworms in soil, biogenic pyromorphite formation may influence Pb mobility and cycling in soils. Lead species in paints and solders. The primary mode of lead exposure in children is through oral ingestion of Pb-containing products, such as lead-based paint and solder. Lead has been used as a pigment since ancient times and has been added to paint in a variety of forms. Perhaps the most common pigment is so-called “lead white,” which contains a lead carbonate phase [2PbCO3·Pb(OH)2]. The popularity of lead white is due to its superior opacity. Other commonly used lead-based pigments include ‘red lead’ (Pb3O4 or 2PbO·PbO2), lead-tin yellow (Pb2SnO4), and Naples yellow [Pb(SbO3)2 or Pb(SbO4)2]. The amount of lead in these pigments may be very high, commonly more than 38% of the dried weight of the paint. In solders, lead is usually present as an alloy with tin. Electronic solders are usually 60% tin and 40% lead by weight in order to produce a near-eutectic mixture. Plumbing solders typically contain a higher proportion of lead. Corrosion of lead-containing plumbing materials in water distribution systems can release lead to drinking water, but the speciation of the lead in the water can vary dependent on municipal water chemistry. Renner (2004) highlights some of the complexities in evaluating the high levels of lead found in drinking water from some areas in Washington, D.C. In most natural waters, lead occurs in the Pb(II) oxidation state. Municipal drinking water, however, tends to be highly oxidizing due to the addition of significant amounts of chlorine. These oxidizing conditions promote the stability of Pb(IV) species such as PbO2 (Fig. 15), which has been observed as a common scaling product in municipal water pipes in Washington, D.C. These scales remain stable and insoluble as long as the water remains oxidizing. However, many municipalities have switched from the use of chlorine to chloramine to comply with the EPA’s 1998 Disinfection Byproducts Rule, which restricts disinfection byproducts in water. Chloramine use lowers the oxidizing potential of drinking water, so that PbO2 scale may become soluble.

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Lead in the body Lead exposure levels. The toxicological effects of lead exposure have been well publicized and are generally familiar to most readers. Humans are often exposed to lead levels above that naturally occurring in soil or dust, and the most common contaminant sources and species have been discussed above. A 1999 Swedish study (Baecklund et al. 1999) characterized lead blood levels (PbB) in 176 men and 248 women, aged 49–92 years. Blood lead levels ranged from 0.56 to 15 µg Pb/dL (median 2.7 µg Pb/dL). The concentration of lead in the blood (PbB) is the most commonly used metric of absorbed dose for lead. The average PbB in the U.S. is ~2 µg/dL (0.1 µM or 20 ppb). The U.S. Centers for Disease Control (CDC) estimates that 890,000 children in the U.S. under 6 years of age may still have unsafe PbB and has established a 10 µg/dL threshold for elevated blood-lead level (CDC 1985). Although elevated lead exposure is generally harmful, the greatest health risks from lead exposure are to preschool-age children and pregnant women and their fetuses. The most profound effects of lead exposure are neurological. Lead affects virtually every neurotransmitter system in the brain, and lead exposure has been linked to schizophrenia. Ancient Romans used lead acetate as a sweetener in wine, which has been thought to have caused a high incidence of dementia. Other medical conditions symptomatic of lead exposure include nephropathy (kidney damage), acute abdominal pain (lead colic or painter’s colic), anemia, bluish discoloration of gums (gingival deposition of lead sulfide, a.k.a. Burtonian lines), gout (Saturnine gout), cardiac toxicity, endocrine effects, reproductive damage, and miscarriage. The EPA has classified elemental lead and inorganic lead compounds as probable human carcinogens. Lead speciation and toxicokinetics. The toxicokinetics of lead in the human body have been extensively studied, but the role of speciation in toxicity is not completely understood. Studies have generally demonstrated that both organic and inorganic forms of lead are readily absorbed through inhalation, although factors such as particle size, solubility, and speciation may play important roles in determining lead absorption rates (ATSDR 2005b). Particles with diameters

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Noble gases in geochemisrty and cosmochemistry (Reviews in mineralogy and geochemistry)

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New Views of the Moon (Reviews in mineralogy and geochemistry 60)

Optical Mineralogy: Principles And Practice

Metals and Society: An Introduction to Economic Geology (Springer Geochemistry Mineralogy)

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